Cadmium Contamination and Its Risk Management in Rice Ecosystems

CHAPTER FOUR

Cadmium Contamination and Its Risk Management in Rice Ecosystems Nanthi S. Bolan*,†,#, Tomoyuki Makino‡, Anitha Kunhikrishnan§, ­Pil-Joo Kim¶, Satoru Ishikawa‡, Masaharu Murakami‡, Ravi Naidu*,†, Mary B. Kirkham**

*Centre for Environmental Risk Assessment and Remediation, University of South Australia, Mawson Lakes, Australia †Cooperative Research Centre for Contamination Assessment and Remediation of the Environment, Adelaide, Australia ‡Soil Environmental Division, National Institute for Agro-Environmental Sciences, Tsukuba, Japan §Chemical Safety Division, Department of Agro-Food Safety, National Academy of Agricultural Science, Gyeonggi-do, Republic of Korea ¶Institute of Agriculture and Life Sciences, Gyeongsang National University, Jinju, Republic of Korea **Department of Agronomy, Throckmorton Plant Sciences Center, Kansas State University, Manhattan, KS, USA #Corresponding author: E-mail: [email protected]

Contents 1. Introduction 2. O  rigin and Sources of Cadmium in Paddy Soils 2.1. G  eogenic 2.2. A  nthropogenic 3. B  iogeochemistry of Cadmium in the Environment 3.1. D  istribution and Speciation 3.2. B  iogeochemistry of Cadmium 4. B  ioavailability and Toxicity of Cadmium 4.1. T oxicity to Plants and Microorganisms 4.2. R  isk to Animals and Humans 5. R  isk Management of Cadmium in Rice Ecosystems 5.1. D  ecreasing Cd Inputs to Rice Soils 5.2. W  ater Management to Reduce Cd Bioavailability 5.3. L ow Cd-Accumulating Rice Cultivars 5.3.1. G  enotypic Variation in Grain Cd Concentration in Rice 5.3.2. Physiological and Genetic Mechanisms 5.3.3. Breeding of Low Cd-Accumulating Cultivars

5.4. S oil Dressing 5.4.1. S imple Soil Dressing 5.4.2. Soil Removal Followed by New Soil Dressing 5.4.3. I n situ Placement of Polluted Soils © 2013 Elsevier Inc. Advances in Agronomy, Volume 119 ISSN 0065-2113, http://dx.doi.org/10.1016/B978-0-12-407247-3.00004-4 All rights reserved.

184 187 188 190 200 200 203 214 215 217 222 223 224 227 227 228 230

231 232 232 233

183

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5.5. P  hytoremediation 5.5.1. N  ecessary Conditions for Phytoextraction 5.5.2. Plant Selection for Phytoextraction 5.5.3. Phytoextraction by High Cd-accumulating Rice

5.6. S oil Washing 5.6.1. S election of Washing Chemicals 5.6.2. On-site Soil Washing (Soil Flushing) in Paddy Fields

5.7. Integrated Risk Management 6. S ummary and Future Research Needs Acknowledgments References

234 234 235 235

243 243 249

250 253 255 255

Abstract Cadmium (Cd) has been identified as one of the major heavy metals reaching the food chain through various geogenic and anthropogenic activities. In many East and South Asian countries including Japan, Bangladesh, Indonesia, and Korea, Cd accumulation in rice (Oryza sativa L.) ecosystems and its subsequent transfer to the human food chain is a major environmental issue. Rice soils in these countries have been affected by Cd accumulation derived from fertilizer and manure application, mine tailings, and refining plants. Excessive intake of Cd into the human body is detrimental to human health, causing serious illnesses such as itai-itai disease. To ensure the safety of foods, the concentrations of Cd in staple crops should be below a standard value; this applies particularly to rice because 34–50% of the Cd intake by people in many Asian countries has been derived from rice. Therefore, development of remediation methods for Cd-contaminated rice soils has become an urgent task to ensure food safety. This chapter provides an overview of the various sources of Cd in rice ecosystems and the biogeochemical processes that regulate Cd bioavailability to organisms, including microbes, plants, animals, and humans. Because of the complexity involved in dealing with Cd in rice ecosystems, exacerbated by the Cd source, site characteristics, and the nature of water management strategies, we have attempted to describe an “integrated” approach that employs a combination of remediation technologies, with the aim of securing methods that are economically and technologically viable.

1. INTRODUCTION Heavy metals reach soils through natural pedogenic (or geogenic) processes and anthropogenic activities. Often the concentrations of heavy metals released into the soil system by pedogenic processes are low and are largely related to the origin and nature of the parent material. Anthropogenic activities primarily associated with agricultural and mining activities, industrial processes, manufacturing, and the disposal of domestic and

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industrial waste materials are the major sources of metal enrichment in soils. Unlike pedogenic input, metals added through anthropogenic activities often have high bioavailability (Adriano, 2001; Naidu and Bolan, 2008). Cadmium (Cd) has been identified as one of the major heavy metals reaching the food chain through various activities (Kirkham, 2006; ­Loganathan et al., 2012; Naidu et al., 1997). For example, in New Zealand and Australia, Cd has been identified as the most common heavy metal reaching the food chain mainly through animal transfer in pastoral agriculture (Loganathan et al., 2012; McLaughlin et al., 1996). Similarly, in many East and South Asian countries including Japan, Bangladesh, Indonesia, and Korea, Cd accumulation in rice ecosystems and its subsequent transfer to the human food chain is a major environmental issue (Kawada and Suzuki, 1998; Simmons et al., 2008) (Tables 4.1 and 4.2). In Australia (Mann et al., 2002; McLaughlin et al., 1996; Williams and David, 1976) and New ­Zealand (Longhurst et al., 2004; Roberts et al., 1994), most of the Cd that has accumulated in the topsoil has been derived from impurities in phosphate (P) fertilizers added during normal farming practice. The paddy soils in many countries have been affected by Cd derived from not only fertilizer application but also mine tailings and refining plants (Luo et al., 2009; Zarcinas et al., 2004). Rice is one of the most widely consumed staple cereal foods in the world constituting about 89% of the diet of people in Asian countries (Chandler and Colo, 1979; Papademetriou, 2000). In these countries, rice is a major source of Cd intake by humans. Excessive intake of Cd into the human body is detrimental to human health, causing serious illnesses such as itai-itai Table 4.1  Selected References on Cadmium Content in Paddy Soils Country Soil Type Concentration (mg kg−1) References

Japan Japan Japan Indonesia Indonesia Indonesia China Macedonia Thailand China Italy Korea

Fluvisols Cambisol Andosols Histosols Luvisols Acrisols Gleysols Acidic soil – Red soil Aquept Long-term ­fertilized soils

0.40 0.60 0.40 0.04 0.09 0.15 0.09 0.1–6.4 0.5–284 2.01–29.68 0.96 1.08–1.16

Herawati et al. (2000) Herawati et al. (2000) Herawati et al. (2000) Herawati et al. (2000) Herawati et al. (2000) Herawati et al. (2000) Herawati et al. (2000) Rogan et al. (2009) Simmons et al. (2005) Yang et al. (2006) Cattani et al. (2008) Jung et al. (2004)

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Table 4.2  Selected References on Cadmium Content in Various Parts of the Rice Plant Concentration Country Element Location (mg kg−1) References

Macedonia 19 countries Malaysia

Thailand China

Japan Taiwan Iran Korea

Grain Grain Grain Husk Leaf Stem Root Stem Leaf Grain Root Straw Stalk Hull Grain with hull Grain without hull Grain Grain Grain

0.005–0.31 0.0008–0.21 0.18 ± 0.028 0.183 ± 0.022 0.203 ± 0.023 0.239 ± 0.386 0.19 ± 0.028 0.38–22.0 0.05–13.5 0.02–5.0 1.79–23.47 0.50–6.24 0.95–5.86 0.55–0.81 0.35–1.15 0.03–0.46 0.15–0.16 0.01 ± 0.04 0.12–0.18

Grain Stem and straw Root

0.013 ± 0.003 0.139 ± 0.001 1.627 ± 0.0.04

Rogan et al. (2009) Watanabe et al. (1989) Yap et al. (2009)

Simmons et al. (2005) Yang et al. (2006)

Nogawa et al. (1983) Lin et al. (2004) Khaniki and Zazoli (2005); Zazoli et al. (2006) Jung et al. (2004)

disease (Nogawa and Kido, 1996; Ogawa et al., 2004). Current Japanese regulations have designated certain paddy fields, which have produced rice grains containing more than 0.4  mg  kg−1 of Cd as “contaminated paddy fields.” Furthermore, the Codex Alimentarius Commission of the United Nations Food and Agriculture Organization (FAO) and the World Health Organization (WHO) recently proposed a new international standard for Cd concentrations in polished rice, 0.4 mg Cd kg−1 (Codex Alimentarius Commission (Codex); Codex, 2005). Therefore, development of remediation methods for Cd-contaminated soils has become an urgent task to ensure food safety. Unlike organic contaminants, most metals do not undergo microbial or chemical degradation, and the total concentration of these metals in soils persists for a long time after their introduction (Adriano, 2001).With greater public awareness of the implications of contaminated soils on human and animal health, there has been increasing interest in the development of technologies

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to remediate contaminated sites. For diffused distribution of metals such as fertilizer-derived Cd input on soils, remediation options generally include amelioration of soils to minimize the metal bioavailability. Bioavailability can be minimized through chemical and biological immobilization of metals using a range of inorganic compounds including lime, P compounds and organic amendments (Bolan and Duraisamy, 2003; Kumpiene et  al., 2007; Park et al., 2011). Reducing metal availability and maximizing plant growth through inactivation may also prove to be an effective method of in situ soil remediation on industrial, urban, smelting, and mining sites. The more localized metal contamination found in urban environments, such as chromium (Cr) contamination in timber treatment plants, is remediated by mobilization processes that include phytoremediation and chemical washing. This chapter focuses on the various sources of Cd in rice ecosystems and the biogeochemical processes that regulate Cd bioavailability to organisms, including microbes, plants, animals, and humans. After laying down the fundamental mechanisms and factors regulating Cd bioavailability, we then assemble the various physical, chemical, and biological mitigative methods that have been demonstrated, highlighting their special strengths and potential for more effective and economical widespread applications in rice fields. Because of the complexity involved in dealing with Cd in rice ecosystems, exacerbated by the Cd source, site characteristics, and the nature of water management strategies, no one remedial technology might suffice. ­Therefore, we have attempted to describe an “integrated” approach of employing a combination of technologies depending on extenuating circumstance, with the aim of securing methods that are, economically and technologically viable. Future research needs, especially in the area of Cd bioavailability and remediation strategies, are identified.

2. ORIGIN AND SOURCES OF CADMIUM IN PADDY SOILS Trace elements include both biologically essential [e.g. copper (Cu), Cr, and zinc (Zn)] and nonessential [e.g. Cd, lead (Pb) and mercury (Hg)] elements (Sparks, 2003). The essential elements (for plant, animal or human nutrition) are required in low concentrations and hence are known as “micro nutrients.” The nonessential trace elements are phytotoxic and/or zootoxic even at low concentration and are widely known as “toxic ­elements” or “heavy metals.” Both groups are toxic to plants, animals and/or humans at excessive concentrations (Adriano, 2001; Alloway, 1990).

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Most trace elements reach the soil environment through both pedogenic and anthropogenic processes. Some of them occur naturally in soil parent materials, chiefly in forms that are not readily bioavailable for plant uptake. Anthropogenic sources include mining and manufacturing activities, and the disposal of domestic and industrial waste materials (Adriano, 2001). Phosphate fertilizers and organic amendments including manures and biosolids are considered to be the major sources of certain trace ­elements [e.g. Cd, Cu, Zn and fluorine (F)] input to soils (Bolan et al., 2003a; Loganathan et al., 2008; Park et al., 2011).

2.1. Geogenic In nature, Cd occurs mainly in association with Zn ores, of which sphalerite (zinc sulfide) forms the main commercial source of Cd. Although Cd occurs in most soil parent materials, its concentration in the common soil-forming rocks such as igneous rocks, sandstones, and limestone is generally low (Bramley, 1990; Loganathan et al., 2012; MacDonald et al., 2005). However, the concentration of Cd in rocks derived from lake sediments and marine black shales is considered to be high. Although none of these rocks occurs to any great extent in most agricultural soils, they can contribute to Cd accumulation, especially in paddy soils derived from these rocks. The presence of black shales containing high concentrations of trace elements including Cd is significant in environmental geochemistry. For instance, the Okchon black shale, which is underlain by the black slates in the central part of the southern Korean Peninsula, provides a typical example of natural geological materials enriched with potentially toxic elements (Kim, 1989).The Okchon Zone of the central part of Korea with an area of about 5100 km2 covers about 5.5% of the total territory of the entire country. Soils derived from these parent materials tend to reflect their extreme geochemical composition (Bowie and Thornton, 1984) and may influence human health by affecting the elemental composition of crop plants (Foth, 1978). In particular, barium (Ba), Cd, molybdenum (Mo), vanadium (V), and uranium (U) in Okchon black shale are highly enriched, and their mean concentrations are significantly higher than those in black slates (Chon et al., 1996; Kim and Thornton, 1993, Table 4.3). Cadmium occurs predominantly as an exchangeable phase in these soils, thereby influencing the high Cd uptake of crop plants. Black shales act as a source of these elements in the rock–soil–plant–human system. Cd concentrations in surface soils are higher in most of the examined

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areas (Chung-Joo, Duk-Pyung, Geum-Kwan, I-Won, Bo-Eun and Chu-Bu) than the world average of 0.4 mg kg−1 Cd quoted by Berrow and Reaves (1984) (Table 4.4; Lee et al., 1998). In the black shale areas, high Cd concentrations are found in residual soils developed from bedrock with high Cd concentrations. High Cd concentrations are also found in alluvial soils in the black slate, black shale, or gray chlorite schist areas. Cadmium concentrations in rice grains are higher in the black shale areas than in the black slate or gray chlorite schist areas. Mean Cd concentrations of rice cultivated in these areas were significantly higher than those in normal rice grown on uncontaminated soils (Masironi et al., 1977; Watanabe et al., 1989;Yoo et al., 1992). Among the six examined areas, the highest mean concentrations Table 4.3  Ranges and Mean Concentrations of Cadmium in Black Shales and Slates from Different Areas in the Okchon Zone, Korea Cd Concentration (mg kg−1) Area

Range

Chung-Joo (N = 7) Duk-Pyung (N = 9) Geum-Kwan (N = 4) I-Won (N = 5) Bo-Eun (N = 5) Chu-Bu (N = 10) Average shale* Average black shale†

0.5–6.5 1.0–36.0 0.4–0.6 0.5–1.0 2.3–3.8 0.5–3.9

Arithmetic Mean

1.4 10.9 0.5 0.6 3.0 1.1 0.3 1.0

*Turekian and Wedepohl (1961). †Vine and Tourtelot (1970). Modified from Lee et al. (1998)

Table 4.4  Cadmium Concentrations (mg kg−1) of Rock, Soil and Rice (Dried Weight Base) Samples in the Duk-Pyung Area in the Okchon Zone, Korea Gray Chlorite Schist, or Black Black Shale Area Slate Area Sample Type

Range

Rock Residual soil Alluvial soil Rice shoot Rice stalk Rice grain

0.4–46 0.3–3.9 0.3–8.3 0.1–2.7 1.0–6.6 0.1–3.5

Arithmetic Geometric Mean Mean Range

Modified from Lee et al. (1998)

6.3 1.3 1.2 0.6 1.7 0.6

2.4 1.0 0.9 0.3 1.0 0.4

0.4–1.7 0.5–1.2 0.3–11 0.1–0.2 0.1–1.7 0.2–0.3

Arithmetic Geometric Mean Mean

0.8 0.8 1.1 0.2 0.5 0.2

0.8 0.7 0.8 0.2 0.3 0.2

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of 0.61 and 1.74 mg kg−1 Cd were found in rice grains and stalks from the Duk-Pyung area, respectively (Table 4.4; Lee et al., 1998).

2.2. Anthropogenic Prior to the industrial revolution, Cd entered agriculture systems at a slow rate as a result of the weathering of rocks and volcanic activity. Now Cd is released into the environment through various industrial processes that include smelting of metals, combustion of coal, oil and wood, and incineration of wastes.The use of P fertilizers and the disposal of sewage sludges and mine tailings directly add Cd to soil. Phosphate compounds contain a suite of trace elements (Bolan et al., 2003a; Loganathan et al., 2008; McLaughlin et al., 1996; Mortvedt, 1996; Syers et al., 1986). Addition of P compounds to soils not only helps to overcome the deficiency of some of the essential trace elements, such as Mo, but may also introduce toxic trace elements, such as Cd and F (Loganathan et al., 2008; McLaughlin et al., 1996; Naidu et al., 1997, T   able 4.5). In this context, Cd contamination of agricultural soils is of particular concern because this trace element reaches the food chain through regular and frequent use of Cd-containing P fertilizer materials, such as single superphosphate (SSP), triple superphosphate (TSP), and phosphate rocks (PRs) (Loganathan et al., 2008). Moreover, presence of Cd in amounts exceeding food guidelines has implications on both local and international trade. Accumulation of Cd in soils through regular fertilizer use has been observed in many countries. For example, in New Zealand and Australia, most of the Cd and F accumulation in pasture soils has been attributed to the use of P fertilizers containing these trace elements (Loganathan et al., 2008). The Cd and F in most P fertilizers originate mainly from the PRs used for manufacturing P fertilizers. It is important to stress that, depending on the source, the PR deposits vary in their Cd content (Bolan et al., 2003a). Thus, manufactured P fertilizers also vary accordingly in their Cd content.The Cd in superphosphates is water soluble and high-analysis P fertilizers, such as TSP and ammonium phosphates, generally contain lower Cd content relative to P (Loganathan et al., 2003). However, depending on the calcium (Ca) content of P fertilizers, the bioavailability of Cd is likely to vary among the P fertilizers.The Ca ion, in addition to competing with Cd for sorption sites, will also reduce the surface negative charge density of soil colloid particles, thereby influencing the bioavailability of Cd (Naidu et al., 1997). Positive relationships between total P and total Cd or F in soils that have received accumulative P fertilizers have often been observed

SSP

188, 376 for 44 years

Pasture

PR, Ca(H2PO4)2

30

Pasture

P

0–780

Mine tailing soil

NC-SSP, NC-PR, Togo-PR, Togo PAPR, Togo-SSP DAP, SSP

126–320

P amended upland paddy soil

450

Horticulture

Phosphate (Na4P2O7·10H2O)

250

Sewage sludge (SS) and P-treated soil

Increased total soil Cd in the 0–75 mm depth from 0.055 to 0.219 mg Cd kg−1 (low P rate) and 0.049–0.432 mg Cd kg−1 (high rates); the rate of Cd accumulation in the topsoil for the high P rate was ­estimated at 7.8 g Cd ha−1 year−1. Cd-containing NPK fertilizers increased Cd ­concentration in ­r yegrass, carrot and ­spinach; Cd uptake decreased with ­increasing pH; low recovery of Cd from PR. Significantly increased radish Cd uptake. Cd uptake by rice grains followed the order of NC-SSP > NC-PR and Togo SSP >  Togo PAPR > Togo PR. The soils had 4–>100 years of ­fertilizer ­history, but the increase in soil Cd was more closely related to Cd input through ­poultry manure than through fertilizers. Cd decreased with incubation time and was reduced by the SS and P additions.

References

Gray et al. (1999a)

He and Singh (1994) Hong et al. (2005) Iretskaya et al. (1998) Jinadasa et al. (1997)

Cadmium Contamination and Its Risk Management in Rice Ecosystems

Table 4.5  Selected References on Sources of Cadmium Input (as a Cocontaminant) to Soils P Loading P Compound* (kg P ha−1) Farming System Observation

Karaca et al. (2002) Continued 191

192

Table 4.5  Selected References on Sources of Cadmium Input (as a Cocontaminant) to Soils—cont’d P Loading P Compound* (kg P ha−1) Farming System Observation

0–400

P-treated paddy soil Pasture

SSP

0, 113, 765

DAP, JPR, NCPR, SSP

0–60 for 10 years

Pasture

HRP, NCPR, PAPR

40–320

Pasture

TSP, K2SO4

250

SSP

188, 376 for 39 years

Biosolid and P-treated soil Pasture

Grain Cd affected but did not exceed the ­permissible concentration. Total Cd increased from 0.10 to 0.4 mg kg−1; ­linear relationship between total P and total Cd in soils. Total Cd concentration increased in the top 120 mm depth; total and plant available Cd was higher in soils treated with high Cd-containing fertilizers. No effect of fertilizer addition on F concentration in clover; whereas fertilizer addition increased Cd concentration in clover. Maximum grain Cd in 20 t ha−1 biosolid +  1/2 P fertilizer treatment. Total Cd in soil ranged from 0.024 to 0.14 mg kg−1 in the unfertilized soils and from 0.085 to 0.34 Cd kg−1 in the Pfertilized soils; obtained a significant correlation between total soil P and Cd.

Lavres et al. (2011) ­Loganathan et al. (1995) ­Loganathan et al. (1996); Loganathan and Hedley (1997) ­McLaughlin et al. (1997) Mousavi et al. (2010) Roberts et al. (1994)

Nanthi S. Bolan et al.

SSP

References

0–188

NPK-fertilized soil

DSP, TSP

0–240

Horticulture

SSP

185–4000

Pasture

DTPA extractable Cd increased with increasing Singh and level of Cd addition; there was no effect of Cd Myhr addition on plant uptake. (1998) Sparrow et al. Potato tuber Cd concentration increased with (1993) increasing amount of Cd applied in P fertilizers; liming did not affect tuber Cd concentration. Both soil and plant Cd levels were higher in the Williams fertilized plots than the control plots; Cd uptake and David decreased with increasing pH (1976)

*DAP—diammonium phosphate, DSP—double superphosphate, FSP—fused superphosphate; JPR—Jordan phosphate rock, NCPR—North Carolina phosphate rock, NCSSP—North Carolina single superphosphate; HRP—Hamrawein phosphate rock; SSP—single superphosphate, TSP—triple superphosphate; PR—phosphate rock; PAPR—partially acidulated phosphate rock.

Cadmium Contamination and Its Risk Management in Rice Ecosystems

NPK fertilizers, PR

193

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(Fig. 4.1; Loganathan et al., 2003). Comparison between unfertilized native and fertilized agricultural soils has often been employed to scrutinize the link between soil contamination and agricultural practices. For example, ­Roberts et al. (1994), Gray et al. (1999a) and Loganathan et al. (1995) obtained significant positive correlations between total P and total Cd in pasture soils in New Zealand receiving long-term inputs of ­superphosphates. Similar results were also obtained for a range of Australian soils (McLaughlin et al., 1996). This is not surprising considering the long history of use in both New Zealand and Australia of superphosphates manufactured from PRs containing high levels of Cd (Rothbaum et al., 1986; Schipper et al., 2011). A number of studies have reported that Cd is derived primarily from P fertilizers used in paddy soils. The application of Mussoorie PR to cultivated paddy areas in India caused an increase in Cd content of the rice grain (Ramachandran et al., 1998), and in Malaysia, small amounts of Cd in the bioavailable form were found to be present in paddy soils (Khairiah et al., 2009). Iretskaya et al. (1998) found that Cd in PR was not readily 0.5

280

0.4

Total F (mg kg–1)

Total Cd (mg kg–1)

240

0.3

0.2

200

160 0.1

120

0 0

400

800

Total P (mg kg–1)

1200

0

400

800

1200

Total P (mg kg–1)

Figure 4.1  Relationship between total P and total Cd or F in soils [●, control; ▲, SSP (single superphosphate); ◆, JPR (Jordan phosphate rock); □, NCPR (North Carolina phosphate rock); ◇, DAP (Diammonium phosphate)] (Loganathan et al., 1999).

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available to rice within few months following its addition to limed soils. However, they noticed that the availability increased markedly in the second year. While the exact mechanism was unclear, they presumed that an increased dissolution of PR, particularly in the acidic soil environments, occurred. Jiaka et al. (2009) studied the effects of different P fertilizers on yields and Cd uptake by paddy rice. Among the three P fertilizer treatments, they observed that amounts of Cd uptake by paddy rice were closely correlated with the amounts of ammonium (NH4) contained in the P fertilizers. The two NH4-bearing P fertilizers [(NH4)2HPO4 and NH4H2PO4] significantly promoted Cd uptake by rice compared to Ca(H2PO4)2; also, (NH4)2HPO4 increased Cd uptake by rice more than NH4H2PO4. This may be attributed to the acidification caused by NH4 fertilizers, thereby resulting in the enhanced mobilization and bioavailability of Cd. In another study, Jamil et al. (2011) evaluated the heavy metal contamination including Cd in slightly acidic cultivated paddy areas, exposed to fluctuating redox conditions. Their results revealed that high amounts of bioavailable Cd in paddy soils might be attributed to the soil type, redox condition, and repeated application of fertilizers in the study areas. The study found that all of the fertilizers commonly used for rice cultivation contained a certain amount of Cd. Although present in small quantities, the investigators warned that Cd could accumulate in soils after several years of application of fertilizers. Recently, Lavres et al. (2011) investigated the P uptake by upland rice and subsequent accumulation in grains from superphosphate fertilizers produced with sulfuric acid treatments of Brazilian PRs. Another objective of their study was to determine the translocation of Cd from P fertilizers to rice grains. Their results revealed that although the grain Cd concentration was significantly affected by P rates, it did not exceed the permissible concentration for human consumption, according to Brazilian food legislation. Conventionally, the term, biosolids (also called “sewage sludge”), refers to the final product derived from the biological treatment of municipal sewage wastewaters. However, recently, the terminology connotates a more inclusive definition to include also livestock manure and other organic wastes, especially when they are composted in the presence of sewage sludge (e.g. greenwaste compost). Most metals, including Cd in biosolids, originate primarily from the contamination of these wastes with industrial wastewater.Traditionally, biosolids are viewed as one of the major sources of metal accumulation in soils, and a large volume of work has been carried to examine the mobilization and bioavailability of biosolids-borne metals in

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soil (Bolan et al., 2003b; Haynes et al., 2009; Park et al., 2011). Advances in the treatment of sewage water and isolation of industrial wastewater in the sewage treatment plants have resulted in a steady decline in the metal content of biosolids (Esmaeily, 2002). Furthermore, stabilization using alkaline materials has resulted in the immobilization of metals in biosolids (Basta et  al., 2001; Bolan et  al., 2003b). Composted alkaline-stabilized biosolids that are low in total and/or bioavailable metal content (known in the USA as “exceptional quality” biosolids; Basta and Sloan, 1999) can be used as an effective sink for reducing the bioavailability of metals in contaminated soils and sediments (Park et al., 2011). However, in most developing countries, biosolids are still considered as a major source of heavy metal input to soils (Haynes et al., 2009). Although many countries have formulated threshold levels for Cd and other heavy metal accumulation in soils due to the use of municipal sewage sludge, such limits have not been established for fertilizer use. Based on the regulatory threshold level for sewage application (3 mg Cd kg−1 soil), the number of years required that would exceed this threshold limit in soil through addition of various sources of Cd including P fertilizer and biosolids is presented in Table 4.6. This indicates that although fertilizer addition represents the major source of Cd input to soils, at the normal annual rate of fertilizer input (20–40 kg P ha−1) to soils, the rate of Cd Table 4.6  Phosphorus (P) and Cadmium (Cd) Contents in Various Phosphate Fertilizers and Organic Amendments, and the Estimated Number of Years Required to Exceed the Threshold Concentration of Cd (3 mg Cd kg−1) in Rice Soils due to their Application Concentration Years Required to Exceed the P (g kg−1) Phosphate Fertilizer Cd (mg kg−1) Threshold Limit*

Single superphosphate Triple superphosphate Diammonium phosphate North Carolina phosphate rock Sechura phosphate rock Egyptian phosphate rock Gafsa phosphate rock Farm yard manure Biosolid Poultry manure Mushroom compost

98 190 200 132 131 130 134 7.5 8.5 17.8 5.3

32 70 10 54 12 10 70 7.6 32 7.5 3.1

149–298 132–264 975–1950 119–238 532–1064 633–1267 93–186 48–96 12–25 115–231 84–166

*At an annual phosphorus application rate of 20–40 kg P ha−1 for rice soils (bulk density of rice soil = 1300 kg m−3 to a depth of 5 cm)

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accumulation appears to be very slow; however, continuous addition of biosolids as a carbon and nutrient source could accelerate the accumulation of Cd in soils. Anthropogenic activities such as mining and smelting of metal ores have increased the occurrence of heavy metal contamination of soil and water sources. Specifically, opencast mining activities have a serious environmental impact on soils and water streams and have generated millions of tons of sulfide-rich tailings (Bhattacharya et al., 2006). Moreover, acidic drainage resulting from the oxidation of sulfides from metalliferous mine spoils leads to the leaching of large quantities of metals including iron (Fe2+), manganese (Mn2+), Pb2+, Cu2+, Zn2+, etc. (Vega et al., 2006). Thus, metal contamination and acid-mine drainage are major environmental concerns where waste materials containing metal-rich sulfides from mining activity have been stored or abandoned (Concas et al., 2006). In general, Cd concentrations in rice and vegetables in the dense mining areas were found to be remarkably higher than those in areas with less mining (Ok et  al., 2011; Zhai et  al., 2008). Long-term Cd exposure by regular consumption of rice and vegetables posed potential health problems to residents in the vicinity of mines. For example, despite the restoration schemes currently in operation, over 1000 abandoned metal mines are still present in South Korea, posing a potential risk to humans and to the ecosystems (Kim et al., 2005). Large amounts of mine wastes including tailings have been left without proper environmental treatment, thus becoming an important point source of toxic elements such as arsenic (As), Cd, Cu, Pb, and Zn in the environment. These materials are dispersed down slope by surface erosion, wind action, and effluent draining from the mine wastes contaminating the low-lying arable lands. In South Korea, 21% of arable soil near mining and industrial areas was found to be contaminated by heavy metals (NIAST, 1997). For example, the Daduk gold (Au)–silver (Ag)–Pb–Zn mine, which is located in the middle part of Korea, is a major source of heavy metal contamination of arable soils. This mine was one of the largest Au–Ag–Pb–Zn mines in Korea. The mine ceased production in 1984 and large amounts of mine wastes have been left without proper environmental treatment. The mine tailings contained high concentration of heavy metals, and they were dispersed down slope by erosion and effluent draining into low-lying land, mainly used for paddy cultivation. Elevated levels of Cd were found in soils sampled in paddy fields and the forest area (Table 4.7). According to the Korean Soil Environmental Conservation Act, soils containing more than

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Table 4.7  Concentration of Cadmium in Mine Tailings and Surface Soils Collected from the Abandoned Daduk Au–Ag–Pb–Zn Mining Area 0.1 N HCl Extractable Total Cd (mg kg−1) Cd (mg kg−1) Sample Type

Tailings and Contaminated area soils Tailing (n = 12) Paddy (n = 34) Forest (n = 8) Control area Paddy (n = 24) Forest (n = 4) Rice plants Contaminated area Grain (n = 8) Stalks and leaves (n = 20) Control area Grain (n = 6) Stalks and leaves (n = 10)

Arithmetic Mean Range

Arithmetic Mean Range

8.57 1.78 1.30

1.56–36.0 0.40–4.76 0.80–2.20

5.67 0.78 0.38

0.05–38.4 0.10–2.49 0.13–1.21

1.18 0.77

0.48–4.24 0.64–0.96

0.35 0.09

0.07–2.5 0.07–0.09

0.15 0.80

0.03–0.65 0.11–5.85

0.09 0.64

0.01–0.24 0.12–2.95

Modified from Lee et al. (2001)

1.5 kg−1 Cd extracted by 0.1 N HCl solution and 4.0 mg kg−1 Cd in total need to be continuously monitored and not used for agricultural purposes, respectively. In urban and semiurban areas throughout Asia, paddy fields for rice production are often close to industrial sites that discharge part of their chemical waste into irrigation channels used for flooding paddy fields (Lin, 2002; Wang et  al., 2007; Plate 4.1). Wastewater discharged from metal plating factories (Higurashi et al., 1976), coating and paint factories (Masui et al., 1971), and electronics and home appliance manufacturing plants (Asami, 1974; Asami et al., 1984; Matsuzaki et al., 1987) is also an important source of Cd contamination.This wastewater often becomes mixed with irrigation water and, thus, contaminates paddy fields. In particular, the metal-smelting process is regarded as the most important Cd contamination source to paddy soils. For example, Cu and Zn mining was the main activity at many nonferrous metal mines in Asian countries like China, Japan, and Korea. At these sites, Cd present in Cu and Zn ores was removed during the concentration or smelting processes and released into the environment with the wastewater and smoke. This

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Plate 4.1  Paddy fields in front of electric factory, leading to heavy-metal ­contamination. For color version of this figure, the reader is referred to the online ­version of this book.

discharged Cd was then carried to agricultural land by water and wind, causing soil contamination (Asami, 1972; Fujimoto and Yamashita, 1976). A Zn smelting factory located in the eastern part of the Korean peninsula is a typical example of Cd contamination resulting from the smelting process. This is the third largest Zn smelting facility in the world. This factory, founded in the 1970s, produces 280,000 tons of Zn, 450,000 tons of sulfuric acid, 1700 tons of Cu, and 900 tons of Cd per year. However, there have been issues reported previously concerning the ill-health effects of heavy metals on exposed workers, living communities, and the environment. Specifically, about 20 ha of arable land near the factory cultivated with different crops were reported to be contaminated by Cd and Zn. The average values for 0.1 N HCl extractable Cd and Zn concentration were 1.7 ± 0.7 and 407  ±  143  mg  kg−1, respectively. About 65% (Cd) and 80% (Zn) of the total sampling sites gave values higher than the warning level (0.1 N HCl extractable Cd 1.5 mg kg−1; Hong et al., 2009). Recently there has been an exponential increase in the production of nickel (Ni)–Cd batteries, especially after the introduction of hybrid car production. For example, in Japan, the domestic production of Cd was over 2000 tons while about 4000 tons of Cd were imported in 2000. Of this amount, the production of Ni–Cd batteries accounted for more than 90% at about 5400–5500 tons. Large numbers of Ni–Cd batteries were distributed domestically and were also exported. Because the recycling rate for batteries is low, it is estimated that more than 1000 tons of Cd may pollute the Japanese environment every year. However, there is no quantitative investigative data on Cd contamination of arable soil by Ni–Cd battery production. In addition,

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the illegal dumping of home electronics products is increasing every year and has become a social problem. Although the number of dumped Ni–Cd batteries and the products using them have not been investigated, it is thought that a considerable amount has been dumped in wooded areas and deserted fields near agricultural land, so that there is the risk of Cd leaking out from illegally dumped Ni–Cd batteries and contaminating agricultural soil.

3. BIOGEOCHEMISTRY OF CADMIUM IN THE ENVIRONMENT 3.1. Distribution and Speciation In soils, Cd occurs in various forms which include free ions in solution, ­soluble and insoluble inorganic and organometallic complexes, ions absorbed onto Fe, aluminum (Al) and Mn hydrous oxides, precipitates such as sulfides, phosphates, and carbonates, and minerals, primarily biotite and riebeckite (Adriano, 2001; Peterson and Alloway, 1979). Understanding the factors controlling Cd speciation and bioavailability in flooded, drained, and alternately flooded/drained paddy soil will be crucial to developing and implementing best management practices needed for productive agricultural areas. Unlike upland soil, lowland rice paddy soil undergoes a flooding and draining cycle, which can change soil conditions into anaerobic and aerobic states and modify the biological and chemical properties of the soil, especially pH and redox potential (Eh) (Kögel-Knabner et al., 2010). These modifications of soil properties may affect Cd speciation present in the ­Cd-contaminated paddy soils. Waterlogging of paddy soils contributes to an increased pH of acid soils and a decrease in pH of alkaline soils. Therefore, the pH tends to converge to neutral, whether the initial soil was acidic or alkaline. Increased pH in acid soils may result in more negative charges on soil clay colloids and organic matter (OM) surfaces, and, therefore, decreases the exchangeable heavy metals by immobilization (Yuan and Lavkulich, 1997). Lim et al. (2002) investigated the changes of speciation of Cd in a tropical coastal clay soil at various pH values at different times and found that the changes of Cd in the exchangeable fraction were pH-dependent. Generally, the change was small under acidic conditions and larger decreases occurred at pH 7. By monitoring the process of heavy metals on iron oxide (α-FeOOH), ­Martínez and McBride (2001) found that either adsorption or coprecipitation of heavy metals with ferrihydrite was pH-dependent. They also found that an increasing pH and incubation time resulted in an increase of

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adsorption and coprecipitation of heavy metals. Zheng and Zhang (2011) suggested that increased concentrations of an OM-bound Cd fraction in flooded soil was probably due to metal–organic complex formation, which has a higher magnitude in waterlogged soils, because lower values of Eh and higher values of pH favored the formation of metal–organic complexes (Gambrell, 1994) and microbial immobilization (Haldar and Mandal, 1979). In waterlogged paddy soils, biological and microbiological activities combined with limited oxygen diffusion causes oxygen depletion, thus resulting in reducing conditions (Kögel-Knabner et al., 2010). In this situation, there is an observable change where a decrement of Eh is followed by an associated increase of pH toward neutrality (Narteh and Sahrawat, 1999). Moreover, the intensity of reduction is higher in the presence of OM due to its oxidizability, and soil components are reduced by anaerobic microbial respiration (Conrad and Frenzel, 2002; Ponnamperuma, 1972). In waterlogged soils, reducing conditions would cause the oxides of Fe and Mn in soil solid phases to be reduced and dissolved (Iu et al., 1981; Ma and Dong, 2004). Chuan et al. (1996) pointed that the pH-dependent metal ­adsorption reaction and the dissolution of Fe–Mn oxyhydroxides under reducing conditions were the mechanisms controlling heavy-metal mobility in acidic soils. Reduced Fe and Mn via hydrolysis and oxidation precipitate as highly amorphous hydrous oxides that have a strong sorption capacity for trace elements (Shuman, 1976). Kashem and Singh (2004) also observed that the breakdown of Fe and Mn oxides caused by waterlogging provided, on one hand, surfaces with high adsorbing capacity for Cd and Zn but, on the other hand, increased concentrations of Fe and Mn in the mobile fraction. Kashem and Singh (2001) suggested that adsorption of trace metals on Fe–Mn oxyhydroxide fractions was the major mechanism of their solubility reduction in submerged conditions. Zheng and Zhang (2011) demonstrated that Cd added to paddy soil in a soluble form under flooding moisture regimes was transformed slowly and consistently from the exchangeable fraction to more stable fractions (Fe–Mn oxide- and OM-bound), leading to the decrease of metal mobility in paddy soil. They suggested that the paddy soil under a flooding regime had higher metal reactivity resulting in more complete transformation of metals to stable fractions, which might be related to increased pH, precipitation of the metals with sulfides, and higher concentrations of amorphous Fe oxides under submerged condition. Cadmium is associated with several mineral phases during flooding periods in paddy soils, including carbonates, humic acid, ferrihydrite, kaolinite, and cadmium sulfide (CdS) (Khaokaew et al., 2011). In general,

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Cd carbonates (CdCO3) and Cd–humic complexes are the major species in lowland paddy soils, while a small amount of CdS was found after the soils were flooded for longer periods (Khaokaew et  al., 2011). Insoluble sulfide forms of metals would be generated in reductive conditions driven by flooding, which could be one of the reasons why heavy metals exhibit low mobility in waterlogged soil. Most metal sulfides are highly insoluble (Sposito, 1994), and under the indirect effects of flooding conditions (low Eh), sulfate ions are reduced to the sulfide form, which might form a complex with heavy metals and immobilize them as sulfide salts (de Livera et al., 2011; Gambrell and Patrick, 1988). Calmano et  al. (1993) reported that generally the easily and moderately reducible fractions increase during oxidation while the sulfidic fraction decreases. Humic acids enhance Cd sorption to hematite, goethite, or kaolinite (Arias et al., 2002; Davis and Bhatnagar, 1995; Lai et al., 2002). Humic acids can directly bind metals via their functional groups, and sorb to oxides and clay minerals, allowing the formation of ternary surface complexes (Arias et al., 2002;Weber et al., 2006). Cadmium bound to humic acids is not stable, especially at lower pH values (Candelaria and Chang, 1997; ­Kunhikrishnan et al., 2012; Zachara et al., 1994). In addition, Cd and Ca can compete for adsorption sites on calcite in alkaline soils since Cd2+ and Ca2+ have similar hydrated ionic radii (Garin et  al., 2003; McBride, 1980; Pickering, 1983; Zachara et al., 1994). Therefore, a mixture of (Cd, Ca)CO3 precipitates can occur at calcite surfaces (Stipp et al., 1992). Mononuclear Cd2+ can chemisorb to the calcite surface at low-Cd concentrations, and/or precipitate to form CdCO3 at high Cd loadings (McBride, 1980; Prieto et al., 2003). Liming causes precipitation of Cd as CdCO3 and a significant decrease of the exchangeable fraction of Cd in contaminated soil (Knox et al., 2001). The formation of Cd carbonates, Cd-kaolinite, and CdS species in paddy soils could limit the release of Cd from soil and Cd uptake by the rice plant. After entering the soil, Cd may be distributed among soil components and associated with them in various forms, which have often been referred to as speciation (Onyatta and Huang, 1999). To characterize the activities of Cd in soil, information about both total concentration and chemical speciation is necessary. However, in order to measure Cd activity in soil and to determine how readily Cd uptake by plants occurs, it is necessary to evaluate the chemical speciation rather than the total Cd content as the former determines the mobility and bioavailability of Cd (Huang et  al., 1988; Kazi et al., 2005). Generally, most speciation studies are carried out using single or sequential extractions using diverse reagents with different

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chemical properties (Arain et al., 2008; Rauret, 1998). Sequential extraction techniques provide a powerful tool for evaluating metal forms (Grzebisz et al., 1997; Shuman, 1985). There are many methods for classifying heavy metal fractionations (Ahnstrom and Parker, 1999; Krishnamurti and Naidu, 2002; Qiao et  al., 2003; Silveira et al., 2006; Tessier et al., 1979). Widely used is Tessier’s fivestep sequential extraction procedure (Tessier et al., 1979), in which heavy metals in soils were categorized in five chemical fractionations including the exchangeable fractionation (F1), carbonate-bound fractionation (F2), Fe–Mn-oxide-bound fractionation (F3), OM-bound fractionation (F4), and residual fractionation (F5) (Silveira et al., 2006;Tessier et al., 1979). F1 is the bioavailable fractionation; F2, F3, and F4 are potentially bioavailable fractionations and F5 is the nonbioavailable fractionation (He et al., 2005; Ma and Rao, 1997). The scheme was developed for sediments but it can also be applied to soils.

3.2. Biogeochemistry of Cadmium The important factors that affect the activity of Cd in soils and its ­availability for plant uptake include soil pH, soil OM content, the metal sorption capacity of the soils, the presence of other micro elements (e.g. Zn) and macro elements (e.g. P), and soil temperature, moisture, and aeration (Bolan et al., 2003a; Chaney and Hornick, 1977; Loganathan et al., 2012). The affinity of Cd for soil surfaces is dependent on the pH and the type of surfaces. The affinity of Cd increases with pH and decreases with concentration of Cd added (Loganathan et al., 2012; Naidu et al., 1994). Adsorption of metals almost invariably decreases with increasing soil acidity (Basta and Tabatabai, 1992; Bolan et al., 1999a; Naidu et al., 1994; Tiller, 1988). Three possible reasons have been advanced for this phenomenon (Naidu et al., 1994). First, in variable-charge soils, a decrease in pH causes a decrease in surface negative charge, lowering cation adsorption. The amount of surface charge acquired through an increase in pH depends on the amount and nature of variable-charge components (Bolan et  al., 1999b). However, Bolan et  al. (2003c) noticed that although there was a positive relationship between increases in pH-induced surface charge through liming and Cd2+ adsorption, only a small fraction of the surface charge was occupied by Cd2+. Second, a decrease in soil pH is likely to decrease hydroxy species of metal cations (MOHn+), which are adsorbed preferentially over mere metal cations (Hodgson et al., 1964). For example, Naidu et al. (1994) observed that CdOH+ species dominate at above pH 8,

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which have greater affinity for adsorption sites than just Cd2+. And third, acidification causes the dissolution of metal compounds, increasing the concentration of metals in soil solution. Liming, as part of normal cultural practices, has often been shown to reduce the concentration of Cd and other metals in the edible parts of a number of crops (Bolan et  al., 2003c; Hong et  al., 2010a; Lee et  al., 2004) (Table 4.8). Addition of other alkaline waste materials such as coal fly ash has also been shown to decrease Cd content of plants (Pourrut et  al., 2011). In these cases, the effect of liming materials in decreasing Cd uptake has been attributed to both decreased mobility of Cd in soils and to competition between Ca2+ and Cd2+ ions on the root surface. In general, Cd uptake by plants decreases with increasing pH. For example, higher Cd concentrations were obtained for lettuce (Lactuca sativa L.) and Swiss chard (Beta vulgaris L.) on acid soils (pH 4.8–5.7) than on calcareous soils (pH 7.4–7.8) (Mahler et al., 1978; Szomolányi and Lehoczky, 2002). Consequently, it is recommended that soil pH be maintained at pH 6.5 or greater in land receiving biosolids containing Cd (Adriano, 2001; Evanylo, 2009). However, it is also possible that in alkaline soils, solubility of Cd can be enhanced due to facilitated complexation of Cd with humic or organic acids (Bolan et  al., 2011; Harter and Naidu, 1995). Thus, the resultant effect of liming on Cd (im)mobilization and subsequent phytoavailability depends on the relative changes in pH and Ca2+ concentration in the soil solution. Ok et al. (2011) conducted laboratory and greenhouse experiments to assess the effects of contaminated rice paddy soils amended with several ameliorants including lime to immobilize Cd and inhibit Cd translocation to rice grain. Sequential extraction analysis revealed that treatment with the ameliorants induced a 50–90% decrease in the bioavailable Cd fractions when compared to the untreated control soil. Their results showed that Cd uptake by rice was decreased by 65% in soils treated with lime compared to the control. But they noticed that ameliorants did not influence rice yield when compared to that of the control. In another pot study, Li et al. (2008) studied the effect of several amendments including limestone on rice growth and uptake of Cd from a Cd-contaminated paddy soil. Their results demonstrated that application of limestone increased grain yield by 12.5–16.5-fold, and decreased Cd concentrations in grain by 50.4%. They suggested that concentrations of Cd in grain and straw were dependent on the available Cd in the soils, and soil-available Cd was significantly affected by the soil pH.

Palygorskite; Sepiolite

Highly polluted mine soil

Ca(OH)2 (8, 15 and 22 Mg ha−1) PR (NCPR)

Limed biosolids (spiked with Cd(NO3)2) Smelter-contaminated soil

CaCO3 (10 g kg−1) KH2PO4, Ca(H2PO4)2

Cd-enriched sewage sludge Variable-charge soil

Ca(OH)2 (8, 15 and 22 Mg ha−1) Ca(OH)2 and CaCO3 (0–1120 kg ha−1) PR waste clay

Sewage sludge Sand Sewage sludge

Maximum Cd sorption capacity at pH 5–6; soluble and readily extractable Cd fraction decreased; 66% and 59% reduction in Cd leaching with palygorskite and sepiolite, respectively. Decreased soil solution Cd and plant uptake of Cd PR decreased the gastrointestinal bioavailability of Cd by 23 and 92% in gastric and intestinal solution, respectively. Decreased Cd phytotoxicity in wheat Cd adsorption increased with increasing level of P, which was attributed to increased negative charge; Cd adsorption was less in the presence of Ca(H2PO4)2 than KH2PO4, which was attributed to increased competition from Ca. Decreased the solution Cd; increased residual fraction and plant uptake of Cd Reduced Cd phytotoxicity

Basta and Sloan (1999) Basta et al. (2001) Bingham et al. (1979) Bolan et al. (1999a; 2003d)

Brallier et al. (1996) Chaney et al. (1977) Gonzalez et al. (1992)

Continued

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DTPA extractable soil Cd and Cd uptake by alfalfa grown on a sludge-amended soil was suppressed by the addition of PR waste clay.

Álvarez-Ayuso and García-Sánchez (2003a; 2003b)

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Table 4.8  Selected References on the Immobilization and Phytoavailability of Cadmium by Various Soil Amendments Soil Amendments Cadmium Source Observation References

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Table 4.8  Selected References on the Immobilization and Phytoavailability of Cadmium by Various Soil Amendments—cont’d Soil Amendments Cadmium Source Observation References

Ca(OH)2

Arable soil-fertilizer Cd

CaCO3 (to pH 7.4)

Arable soil/sewage sludge

CaCO3, Ca(OH)2, CaSO4.2H2O, Oyster shell meal

Decreased Cd in chemical extractants and plant tissue Decreased Cd2+ adsorption

Hooda and Alloway (1996) Hong et al. (2007)

John and Van Laerhoven (1976) Kreutzer (1995) Maier et al. (1997) Mandjiny et al. (1998)

McGowen et al. (2001)

Nanthi S. Bolan et al.

Heavy-metal-contaminated soil Ca(OH)2 was found to be more efficient on reducing soil NH4OAc extractable Cd and plant-Cd concentrations, due to increased net-negative charge of soil induced by pH increase. CaCO3 (17.92 Mg ha−1) Sewage sludge and milorganite Decreased uptake of Cd by plants resulting in Cd attenuation Decreased Cd concentration in soil solution CaMgCO3 (4 Mg ha−1) Forest soil Increased Cd concentration in potato tuber CaCO3 (0–20 Mg ha−1) Arable soil Hydroxyapatite (HA) Cd solution (spiked with HA removed Cd from aqueous solutions Cd(NO3)2) with efficiency higher than 99.5% at pH 5–6. XRD and SEM indicated that Cd is incorporated into the hydroxyapatite structure via diffusion and ion exchange. (NH4)2HPO4 Smelter-contaminated soil P decreased the leaching of Cd through precipitation as metal phosphate; solubility diagram provided evidence for Cd3(PO4)2.

Gray et al. (1999b)

Mine tailings soil

After 30 days of incubation, 0.1 N HCl extractable Cd in soil decreased significantly as a result of an increase in soil pH and the formation of metal hydroxides. COSP was more effective in immobilizing Cd than NOSP. Zerovalent iron (ZVI), Metal-contaminated paddy soil Compared to the control Cd uptake by rice lime, humus, compost was decreased by: ZVI + humus (69%), lime (65%), ZVI + compost (61%), compost (46%), ZVI (42%), and humus (14%). CaCO3 (0–2.5 Mg ha−1) Arable soil Decreased Cd concentration in barley grain K2HPO4 Cd-amended artificial soil Decreased the solubility and bioavailability of Cd to earthworms. XRD indicated Cd3(PO4)2 formation. Hydroxyapatite (HA) Contaminated soil Decreased solution concentration through secondary precipitates rather than sorption by weathered hydroxyapatite grains. Aqueous palygorskite, Cd-amended soil Average amount of Cd released after five sepiolite and calcite desorption steps was 13.8%, 2.2% and 3.6% for the palygorskite, sepiolite and calcite, respectively, indicating that a large portion of Cd was irreversibly retained by the minerals.

Ok et al. (2010)

Ok et al. (2011)

Oliver et al. (1996) Pearson et al. (2000) Seaman et al. (2001) Shirvani et al. (2006)

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Natural oyster shell powder (NOSP) and calcined oyster shell powder (COSP)

Continued 207

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Table 4.8  Selected References on the Immobilization and Phytoavailability of Cadmium by Various Soil Amendments—cont’d Soil Amendments Cadmium Source Observation References

CaCO3

NPK fertilizer

CaCO3 (1–5 g kg−1)

Arable soil

P-rich biosolid

Cd-spiked soil (spiked with Cd(NO3)2)

CaCO3

Pasture soil

Ca(OH)2, Sepiolite

Cd-contaminated paddy field

Application of Ca(OH)2 and sepiolite increased soil pH; NaNO3 and CaCl2 extracted Cd reduced by 61–100% and 52–98%; Cd in rice reduced significantly; sepiolite or sepiolite-lime more effective than lime alone.

Singh and Myhr (1998) Singh et al. (1995) Soon (1981)

Tyler and Olsson (2001) Zhu et al. (2010)

Nanthi S. Bolan et al.

SEM–scanning electron microscope; XRD–X-ray diffraction.

Decreased DTPA and NH4NO3 extractable Cd; increased plant-tissue Cd Decreased DTPA and NH4NO3 extractable Cd, and plant-tissue Cd Cd adsorption increased with increasing pH; about 82–92% of adsorbed Cd was retained by cation exchange and complexing sites; adsorption by cation exchange became more dominant as the amount of Cd in the soil was increased. Decreased Cd concentration in plant tissue

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Liu et al. (2011) examined the effects of lime and other alkaline substances on remediation of Cd-contaminated paddy soil using a pot experiment. The results showed that a single application (2 g kg−1) of lime and other alkaline substances increased soil pH and soil-exchangeable Ca content.The soil-available Cd content was decreased by 37.4% and 33.2% at 30 and 60 days after rice transplantation, respectively. Their study revealed that Cd content in rice root and brown rice was decreased by 24.0% and 26.3%, respectively. Cattani et al. (2008) conducted a field trial during 2003 and 2004 in Italy to study reduction of the uptake of Cd by rice using lime as an amendment. They found that submersion was the main factor decreasing Cd concentration in rice grain, producing maximum concentrations of 0.14 and 0.06 mg kg−1 in 2003 and 2004, respectively. They observed that Cd concentrations were at least two times higher for rice cultivated by irrigation only than under submerged conditions. The addition of lime decreased the Cd concentration of rice by about 25% compared with control under dry conditions.They suggested that lime addition appeared to be a promising technique to reduce the bioavailability of soil Cd and minimize Cd concentrations in the rice. Differences in uptake over the years, with concentrations up to 40% lower in 2004, were explained by differences in transpiration. Their results also demonstrated that the influence of climatic conditions on Cd uptake in plants should not be underestimated. Metals form both inorganic and organic complexes with a range of solutes in soils. For example, a number of studies have indicated that chloride (Cl−) forms a soluble complex with Cd as CdCl+, thereby decreasing the adsorption of Cd onto soil particles (McLaughlin et al., 1996; Naidu et al., 1994, 1997; Weggler-Beaton et al., 2000). Similarly, the formation of aqueous complexes with dissolved organic carbon (DOC) and low-molecular weight organic acids from root exudates is expected to dominate the solution chemistry of certain metals, such as Cd and Cu in the rhizosphere (Bolan et  al., 2011; Fotovat and Naidu, 1997; Krishnamurti et  al., 1996; Kunhikrishnan et  al., 2012). Although the effect of organic amendments such as composted biosolids on Cd2+ adsorption is not consistent, the phytoavailability of Cd has been shown to be greatly inhibited (Bolan et al., 2003b; Park et al., 2011). Addition of composted biosolids increases surface charge, which may be one of the reasons for increasing Cd adsorption, thereby decreasing phytoavailability. Plants exhibit greater tolerance to metals introduced through sewage sludge addition than when they are added as inorganic salts. For example, Chang et  al. (1992) and Logan et  al. (1997) presented data for maize

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(Zea mays L.) and other crops grown on metal-contaminated sludgeamended soils, which revealed an inconsequential change in crop tissue Cd concentrations in response to substantial increases in total Cd loading in soils. The decrease in the phytoavailability of metals in the presence of organic amendments is often attributed to increased complexation of the metal by the organic constituents (Adriano, 2001). However, the presence of phosphates, Al and Fe compounds, and other inorganic minerals in typical municipal sewage sludge is also believed to be responsible for inducing the “plateau effect” in Cd uptake by crops, thereby preventing the increased Cd availability suggested in the “time bomb” hypothesis (Brown et al., 1998). In soils containing large amounts of OM, such as pasture soils and organic manure-amended soils, only a small proportion of the Cd in soil solution remains as free Cd2+ and a large portion is complexed with soluble organic carbon (del Castilho et al., 1993; Sauve et al., 2000). Addition of manure and composted biosolids has been found to increase the complexation of Cd in soils, the extent of which is related to the amount of DOC. Although a wide variety of organic compounds in DOC are involved in the formation of soluble complexes with metals, Zhou and Wong (2001) and del Castilho et al. (1993) observed that low-molecular fractions, such as hydrophilic bases, have a strong affinity to form soluble Cd complexes. Similarly Riffaldi et  al. (1983) obtained significant correlations between Cd sorption and phenolic hydroxyl groups and carboxyl groups of fulvic acids. Fulvic acid, although representing an average 9% of the total organic C content in sludges, plays an essential role in metal retention (Bolan et al., 2011; Senesi and Loffredo, 1999; Senesi and Plaza, 2007). Farmyard manure (FYM), including cow or pig manure, decreases bioavailability of heavy metals in soil and crop plants (Pichtel and Bradway, 2008; Ram and Veerloo, 1985). FYM positively influences crop production (Kaihura et al., 1999), improves soil physical properties (Chen et al., 1996), and can be used to reduce heavy metal hazards in plants (Yassen et al., 2007). Li et al. (2009a) investigated the effects of pig manure on the distribution and accumulation of Cd in a soil–rice system using a pot experiment. Results showed that application of pig manure decreased the concentrations of Cd in rice roots by 35.6%. They observed that pig manure not only decreased uptake of Cd by rice but also restrained the transfer of Cd from the rice root to the stem and grain.The application of amendments increased soil pH and resulted in the reduction of Cd concentrations in soil solutions, which were significantly correlated to the uptake of Cd by the rice stem and grain. In a

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similar study, Kibria et al. (2011) also noticed a significant reduction in the grain Cd content of rice by 27% by FYM. Han et  al. (2011) conducted a glasshouse experiment to examine the Cd concentrations in the aboveground parts of rice after application of pig manure in three soils differing in pH, texture, OM, and CEC. The results demonstrated that soil pH increased with increasing rates of pig manure from 1% to 3%, and the highest rate of manure produced lower Cd concentrations in the grain, husk, and straw on all three soils. They noticed that grain Cd concentrations were lower in soil with the highest pH, but in other soils, it exceeded the guideline value of 0.2 mg kg−1. In another pot study, Li et al. (2009b) applied composted pig manure to soils at rates of 0%, 0.5%, 1.5%, 3.0%, and 5.0% which were equivalent to 0, 0.1, 0.3, 0.6, and 1.0 mg kg−1 of Cd to assess the Cd accumulation by rice plants. Results indicated that Cd concentrations in rice grains were more than the limit of 0.2 mg kg−1 when 0.14 mg kg−1 or more Cd was loaded to Ferralsols by manure application, but it was not more than the limit in Calcaric Cambisols (Table 4.9). They also found that in all treatments, Cd accumulations in rice straw and roots were significantly greater than those in rice grains and only a small portion of Cd absorbed by rice plants was transferred into the grains. With composted biosolids, one of the main concerns about its longterm metal-immobilization efficiency is the potential for metals to mobilize if and when the OM undergoes significant oxidation. Stacey et al. (2001) observed that the release of Cd and Zn from composted biosolids, as the OM in the compost decomposed, depended to a large extent on the composition of the composted biosolids. However, Hyun et al. (1998) obtained no evidence for increased phytoavailability of Cd with the breakdown of Table 4.9  Cadmium Contents in Rice and Soil from Various Soil Types (Herawati et al., 2000) Soil type In Rice (mg kg−1) In Soil* (mg kg−1) In Soil† (mg kg−1)

Andosols Cambisols Fluvisols Gleysols Histosols Luvisols Acrisols Mixed soils

0.0667 0.0494 0.1640 0.0122 0.0182 0.0245 0.0404 0.0366

*Metals extracted by hydrochloric acid. †Metals ashed by nitric acid.

0.0603 0.1484 0.0737 0.0447 0.0122 0.0271 0.0300 0.0814

0.4034 0.6018 0.4037 0.0995 0.0366 0.0900 0.1484 0.4034

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OM in sludge-treated soils. Furthermore, Hettiarachchi et al. (2006) and Li et al. (2001) observed evidence for greater affinity for Cd adsorption by the inorganic components of the composted biosolids-amended soils indicating that the increased adsorption of Cd is independent of the added OM and of a persistent nature. Although the formation of soluble metal–organic complex reduces the phytoavailability of metals, the mobility of the metal may be facilitated greatly in soils receiving alkaline-stabilized composted biosolids because of an increased concentration of soluble metal–organic complexes in solution (Brown et al., 1997; Dinel et al., 2000; Gove et al., 2001). Soluble complexing ligands in biosolids cause certain trace elements to be more mobile than they would be in the absence of organics (Ashworth and Alloway, 2008; Camobreco et al., 1996; Frenkel et al., 1997). Camobreco et al. (1996) found that soluble complexing ligands in biosolids enabled metals with low (Cd, Zn) and high (Cu, Pb) organic affinities to move through undisturbed soil columns. Lamy et al. (1993) also observed DOC facilitation of Cd mobility following sludge application. The behavior of Cd in a mixed sludge–soil system showed that the addition of sludge-soluble OM to the soil led to a decrease of Cd sorption across the pH range between 5 and 7. Antoniadis and Alloway (2002) investigated the leaching of Cd, Ni, and Zn down the profile of sewage sludge-treated soil from packed columns and found that the metals moved significantly down the soil profile to a depth of at least 8 cm at 50 t ha−1. They suggested that DOC increased the mobility of these heavy metals by acting as a ligand in the soil with sewage sludge application.They cautioned that the depth of metal movement could increase at high sludge-application rates and in areas with high precipitation, even where the soil pH does not necessarily encourage metal solubility. For most heavy metals and sludge types, the highest dissolved metal concentrations appear in leachates during or shortly after the sewage sludge amendments (Richards et al., 2000). However, McBride et al. (1999) observed that trace metals including Cd, Zn, Ni, and Cu continued a gradual process of leaching from contaminated subsoil and through the shallow subsoil more than 15 years after sewage sludge application. McBride et al. (1999) observed that 82% of the Cd in percolates collected at the 60 cm depth from a long-term study was largely in complexed form with DOC. In another similar study, Schaecke et  al. (2002) investigated the fate of heavy metal concentrations in soil fractions including Cd after biosolid application at 0–0.25 m depth. They observed that 11 years after the last application, metals supplied with the sludge had moved as far as 50 cm in

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depth. Concentrations of metals in the saturation extract of the sampled soil layers were closely correlated to the concentrations of DOC. Zubillaga and Lavado (2008) examined the accumulation of Cd, Cu, Pb, and Zn in soils throughout 5 years, during and after biosolid application and their movement with depth in a Typic Argiudoll and a Typic Hapludoll. They noticed that only total Cd in the Typic Argiudoll moved with depth, which they attributed to the metal complexation by DOC released from the biosolids. Data from laboratory and glasshouse experiments have clearly demonstrated that P addition enhances the immobilization of Cd in soils, thereby alleviating its phytotoxicity (Bolan et  al., 2003d; Kumpiene et  al., 2007) (Table 4.8). Two reasons could be attributed to P-induced immobilization of Cd in soils: (1) P-induced Cd2+ adsorption; and (2) precipitation of Cd as Cd(OH)2 and Cd3(PO4)2. Several mechanisms can be advanced for P-induced Cd2+ adsorption. These include: (1) increase in pH; (2) increase in surface charge, (3) co-adsorption of P and Cd as an ion pair, and (4) surface complex formation of Cd on the P compound (Bolan et al., 2003d). The effect of P addition on soil pH depends on the buffering capacity of the soil, the nature of P compounds, and the extent of P adsorption (Havlin et  al., 1999). While the addition of nitrogen (N)- and Ca-containing P ­fertilizers generally decreases soil pH due to acidification reactions, addition of other P fertilizers increases soil pH mainly due to the ligand-exchange (i.e. OH− ions) P adsorption reaction. Levi-Minzi and Petruzzelli (1984) observed that P-induced variation in soil pH influenced the solubility of Cd in soils.They noticed that while the effect of P on pH and Cd solubility was less evident in an organic soil with high pH-buffering capacity, the addition of diammonium phosphate (DAP) increased soil pH, thereby reducing the solubility of Cd in a mineral soil with less pH-buffering capacity. A number of studies have shown that specific adsorption of anions increases the net negative charge of variable charge surfaces (Bolan et al., 2003d; Bolland et al., 1977; Kuo, 1986; Lackovic et al., 2003; Naidu et al., 1994, 1997).The amount of surface charge acquired through specific adsorption depends on the nature of anion adsorbed and the pH and electrolyte concentration of the solute. Naidu et  al. (1994) and Bolan et  al. (1999a; 2003d) showed that increasing P addition caused a significant increase in Cd2+ adsorption by soils dominated by variable-charge components. It has been shown that Zn2+, Cd2+, or Cu2+ adsorption by variable-charge components, such as Al and Fe hydrous oxides, can be increased by low or moderate enrichment of oxides with P (Bolland et al., 1977; Hong et al.,

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2008, 2010b; Kuo, 1986), which has been attributed to increases in surface negative charge after P adsorption. While surface complex formation has also been offered for the increased adsorption of Ca2+ or Zn2+ onto P-enriched gibbsite and goethite by ­Helyar et al. (1976) and Bolland et al. (1977), Xu et al. (1994) concluded that coprecipitation was the dominant process in the adsorption of Cd2+ (Eqn (1)) and Zn2+ by hydroxyapatite:

Ca10(PO4)6(OH)2 + xCd2+ → (Cdx,Ca10−x)(PO4)6(OH)2 + xCa2+ (1)

Precipitation as metal phosphates has also been proved to be a main mechanism in immobilizing metals, such as Cd, Pb, and Zn, by P compounds (Table 4.8). McGowen et al. (2001) have examined the immobilization of As, Cd, Pb, and Zn in a smelter-contaminated soil using DAP. Application of high levels of DAP at a rate of 2300 mg P kg−1 was effective in immobilizing Cd, Pb, and Zn in the contaminated soil. Others have also shown that Cd3(PO4)2 can control Cd solubility in P-sufficient soils or soils amended with P (Hong et  al., 2010b; Santillan-Medrano and Jurinak, 1975; Street et al., 1978). However, because the solubility of Cd3(PO4)2 is too high to control the concentration of Cd in soils, it is unlikely this solid phase can play a significant role in the immobilization of Cd (Bolland et  al., 1977; Soon, 1981; Xiong, 1995, Table 4.8).

4. BIOAVAILABILITY AND TOXICITY OF CADMIUM The bioavailability of a chemical in the soil environment has been defined as the fraction of the total element in the interstitial pore water (i.e. soil solution) and soil particles that are available to the receptor organism (Naidu et al., 2008). Considerable controversy exists in the literature relating to bioavailable fraction, including the definition itself and the methods used for its measurement. A more generic definition of bioavailability is the potential for living organisms to take up chemicals from food (i.e. oral) or from the abiotic environment (i.e. external) to the extent that the chemicals may become involved in the metabolism of the organism. More specifically, it refers to the biologically available chemical fraction (or pool) that can be taken up by an organism and can react with its metabolic machinery (Campbell, 1995); or it refers to the fraction of the total chemical that can interact with a biological target (Vangronsveld and ­Cunningham, 1998). In order to be bioavailable, the trace elements have to come in contact with

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the organism (i.e. physical accessibility). Moreover, trace elements need to be in a particular form (i.e. chemical accessibility) to be able to enter biota. In essence, for a trace element to be bioavailable, it will have to be mobile and transported and be in an accessible form to the biota concerned.

4.1. Toxicity to Plants and Microorganisms Although Cd is a nonessential element for both plants and animals, increasing Cd concentration in soil leads to an increase in Cd uptake below the phytotoxicity threshold concentrations. Plant species differ in their ability to extract soil Cd and, in general, weed species are shown to accumulate more Cd. For example, a nationwide survey by Roberts et al. (1994) indicated significantly higher Cd concentration in weeds (0.28 mg Cd kg−1) than grasses and legumes (0.06–0.1  mg  Cd  kg−1). Similarly, in Australia, the highest Cd concentration is obtained in cape weed (1.57 mg Cd kg−1), which is a common component of pastures in Australia and New Zealand. Loganathan et al. (2003), McLaughlin et al. (1997), and Williams and David (1973) have shown that the concentration of Cd in legumes is several times higher (0.155 mg Cd kg−1) than the associated grass (0.031 mg Cd kg−1). The increased concentration of Cd in legumes may be related to the acidifying effect of legumes (Tang et al., 1998). It is well known that Cd concentration in plants vary with plant species (Adriano, 1986; Alloway, 1995; Almas et al., 2006; Black et al., 2011). In ­general, leafy plants tend to accumulate higher metal concentrations than root, grain, or fruit crops (Alloway, 1995). Cadmium accumulation in rice grains cropped on Cd-contaminated soils could increase 6- to 10-fold depending on the genotype of the cultivar used (Arao and Ishikawa, 2006; Yu et al., 2006). Both soil properties which control Cd availability and differences between cultivars are relevant in order to assess risks of Cd in soil in relation to the quality of rice. Although numerous factors contribute to final Cd levels in rice grains, the increase in Cd uptake by rice coincides with an increase in the available Cd pool in soil. Cadmium uptake levels of rice grains could be quite different among cultivars under similar soil conditions (Table 4.10). In general, Cd accumulation in rice grains of Indica species proved to be high compared to Japonica species (Römkens et al., 2009). Hence, Japonica species are more suitable for cropping on soils affected by Cd pollution (He et al., 2006). Irrespective of rice cultivars, the combination of elevated total Cd concentrations in soil, a low pH, and low soil OM content results in an increased availability of Cd in soil, which results in a high uptake of Cd by rice plants (Römkens et al., 2009).

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Table 4.10  Overview of Cadmium Concentrations in Roots and Grains of Indica and Japonica Rice Cultivars Total Cd in Soils (mg kg−1) Cd in Roots (mg kg−1) Cd in Grains (mg kg−1) Cultivar Name

Japonica Japonica Japonica Japonica Japonica Japonica

Tainung no. 70 Taiken no. 8 Tainung no. 72 Kaohsiung no. 143 Taitung no. 30 Tainung Sen no. 20 Tainung no. 71 Tainung no. 67 Kaohsiung Sen Yu no. 1151 Taichung Sen Waxy no. 1 Taichung Sen no. 10 Kaohsiung no. 144

Japonica Japonica Indica Indica Indica Indica

Modified from Römkens et al. (2009)

Median

Range

Median

Range

Median

Range

Ratio Cd-Rice Grains/Roots

0.60 0.61 0.65 0.64 0.59 0.71

0.13–27.8 0.09–18.8 0.11–18.2 0.07–17.4 0.06–23.9 0.08–21.2

8.0 6.6 9.7 8.0 8.8 6.5

0.8–373.4 0.5–181.3 0.4–403.9 0.5–198.5 0.7–213.6 0.6–247.7

0.21 0.23 0.19 0.19 0.18 0.43

0.02–4.6 0.11–6.0 0.02–3.0 0.01–4.5 0.02–3.3 0.02–12.6

0.029 0.037 0.026 0.029 0.025 0.096

0.60 0.66 0.57

0.13–25.9 0.08–26.6 0.09–25.9

7.2 5.4 6.2

0.6–175.7 0.6–139.2 0.5–107.0

0.20 0.19 0.44

0.10–3.7 0.09–3.4 0.23–7.6

0.030 0.032 0.075

0.71

0.14–25.8

6.4

0.5–161.8

0.60

0.25–25.3

0.092

0.70

0.14–22.7

9.6

0.7–266.6

0.37

0.19–29.1

0.061

0.59

0.13–16.8

4.6

0.6–107.7

0.16

0.08–3.7

0.032

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Despite observed differences in Cd levels in grains of Indica versus J­aponica cultivars (Table 4.10), no significant differences in Cd-root levels were observed (Liu et al., 2003; Römkens et al., 2009).This finding suggests that Indica and Japonica cultivars mainly differ in their ability to transfer Cd from the root into the shoot. Indeed, the ratio of Cd in rice grains to that in the root differs substantially between Japonica (0.03) and Indica (0.08), which explains the higher Cd levels in rice grains of Indica cultivars. In general, Cd uptake by rice roots is known to be affected by competitive cations having similar electronic properties such as Ca, magnesium (Mg), and Zn. For example, the effect of Zn on uptake of Cd by roots and rice grain has often been reported (Girling and Peterson, 1981; Hassan et al., 2006), although contrasting effects have been demonstrated depending on the levels of Zn and Cd in the soil and the soil Cd-to-Zn ratio (Dunbar, 2004; Kukier and Chaney, 2002). Liu et al. (2007) found that Zn in solution suppressed Cd uptake by roots but increased the transfer of Cd from root to rice grains.

4.2. Risk to Animals and Humans The tolerable intake for Cd as proposed by FAO/WHO (1972) is 400–500 µg week−1 person−1 or 57–71 µg day−1 person−1 weighing 70 kg. Despite the difference in the amount of dietary Cd intake, however, rice is the leading source of Cd common to many populations in Asia, e.g. approximately ranging from 20% of total dietary Cd intake in the Philippines (Zhang et al., 1998), to 30–40% in Japan (Ikeda et al., 1999; Watanabe et al., 2000a), many parts in mainland China (Zhang et al., 1997), Taiwan (Ikeda et al., 1996), and Thailand (Zhang et  al., 1999), and with a much higher level (53%) in Malaysia (Moon et al., 1996). It may be of interest to stress that wheat (Triticum aestivum L.) and other cereals are additional sources of dietary Cd when people consume such cereals more than rice, as observed in northern China (Watanabe et al., 1998, 2000b). It is, however, worth noting that while rice is the major component of food among the Asian community, many of them (especially females and children) weigh <70 kg. Thus FAO/ WHO tolerable limit for Cd and indeed other potentially toxic substances may not be appropriate for Asian communities. In comparison, dietary Cd intake levels in the 1990s in Europe have been made available in recent years. The levels reported (mostly geometric mean or medians for adult people, unless otherwise specified) include 6.3 µg week−1 person−1 for 1.5–5.3-year-old children in a remote island in the North Sea, Germany (Schrey et al., 2000), approximately 7 µg day−1

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in Duisburg, Germany (Wilhelm et  al., 1995), 9–22  µg  day−1 in Sweden ( Järup et  al., 1998), 11.1  µg  day−1 for students in Spain (Barbera et  al., 1993), 14 µg day−1 in the UK (Ysart et al., 1999), 17.5 µg day−1 in Croatia (­Sapunar-Postruznik et  al., 1996), 19–27  µg  day−1 in Belgium (Van ­Cauwenbergh et al., 2000), and 28 µg day−1 in the Ruhr district, ­Germany (Wilhelm et  al., 2002). The value reported for Canadians (Dabeka and McKenzie, 1995), 13 µg day−1, was also close to these values observed in Europe. The levels as a whole appear to be close to or slightly higher than the levels in Asia (5–15 µg day−1; Ikeda et al., 2000) excluding Japan where the exposure level is much higher (i.e. ∼30 µg day−1). In particular, background exposure of general populations in Japan to Cd has been relatively high (Watanabe et al., 2000a). According to Watanabe et al. (1996), there is a significant difference in geometric mean Cd content in rice samples (1546) collected in 17 areas in the world, but it was the highest in Japan (0.0557 mg kg−1) and Colombia (0.13320 mg kg−1) (Table 4.11). Analysis of data from food duplicate surveys in the mid 1990s on some 600 women throughout the country coupled with air pollution data disclosed that foods were the almost exclusive source of Cd exposure for the general population, whereas the exposures through the respiration route was essentially negligible; rice (paddy rice), the staple cereal for most Japanese, was in fact the leading source of Cd intake via daily foods, accounting for 30–40% in the 1980s and 1990s (Ikeda et al., 1999; Watanabe et al., 2000a), although a substantial decrease in total dietary intake occurred in the mid 1990s as compared with the levels 10–15 years ago (Watanabe et al., 1996, 2000a). The Japanese people used to take in approximately 8.2 and 1.8  µg  Cd  day−1 from rice and wheat, respectively (Shimbo et  al., 2001) (Table 4.12). It is worth noting that the dietary route is responsible for 99% of Table 4.11  Cadmium Contents in Rice Grain from Various Areas (Watanabe et al., 1996) Areas Mean Cd (mg kg−1) Areas Mean Cd (mg kg−1)

Australia China Taiwan Indonesia Japan Korea Thailand Malaysia Philippines

0.00267 0.01554 0.03955 0.02177 0.05570 0.01570 0.01570 0.02774 0.02014

Vietnam Canada Colombia Finland France Italy South Africa Spain USA

0.01850 0.02902 0.13320 0.02580 0.01741 0.03392 0.02582 0.00085 0.00743

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total Cd intake in Japan because Cd concentration in general atmospheric air is low compared to food (Ikeda et al., 2000). The dietary Cd intake by middle-aged Japanese women in the late 1990s was 25.5 µg day−1 (­Watanabe et al., 2000a) (Table 4.13).This amount of Cd intake in Japan appeared to be 2–3 times higher than the levels among other women populations in east and south-east Asia (Ikeda et al., 2000), possibly because Cd contents in the rice are lower there, even though in these places people depend even more heavily on rice as a major source of energy for daily life. Cadmium reaches the food chain also through uptake by grazing animals. For example, in New Zealand, most of the Cd accumulated in grazing animals is derived mainly from the pasture intake (Loganathan et al., 2008). Bramley (1990) has estimated that annually ∼55 and 275 mg Cd is ingested per sheep and cow, respectively, through the intake of herbage. Work in New Zealand has indicated that cattle and sheep may ingest 1–10% and >30%, respectively, of their dry matter intake in the form of soil. In areas where soils contain high Cd, for example, as a result of sewage sludge and fertilizer application, soil ingestion is expected to play a significant role in Cd uptake by farm animals (Loganathan et al., 2003;Wilkinson et al., 2001). Roberts et al. (1994) have shown that even though sheep ingest 36–46 kg soil per year, this only contributes 5–8% of the total Cd intake for the lax and hard grazed flocks, respectively. Ruminants do not have a homeostatic control mechanism for regulating Cd absorption or excretion which is affected by the level of dietary Cd content (Lee et  al., 1996; Loganathan et  al., 2008). Although intestinal uptake of Cd has been estimated to account for over 90% of the Table 4.12  Intake of Cadmium by Japanese via Rice and Wheat in 1998–2000 Rice Wheat Item

Average

Ranges

Average

Ranges

Daily cereal consumption (g day−1) Cadmium content (mg kg−1)* Daily Cd intake (µg day−1)

165

158–178

91

67–104

0.0497

0.0430–0.0701

0.0193

0.0170–0.0212

8.2

7.409–12.506

1.754

1.297–1.993

*Cited from Ministry of Health and Welfare (2000) Revised from Shimbo et al. (2001)

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Table 4.13  Dietary Exposure to Cadmium (Cd), and Cd Concentration in Blood and Urine among Nonsmoking Adult Women in the Cities in Asia Cadmium in Blood (µg L−1)

Urine (mg kg−1 References Creatinine)

Japan (A group of 27 sites) 1977–1981 37.5

3.47

–

1991–1997

25.5

1.90

4.39

Bangkok

7.1

0.41

1.40

China (A group of 4 sites) Korea (A group of 4 sites) Kuala Lumpur

9.9

1.07

2.30

21.2

1.39

2.26

9.0

0.74

1.51

Manila

14.2

0.47

1.21

Tainan

9.7

0.83

1.59

Location

Food (µg day−1)

Watanabe et al. (2000a) Watanabe et al. (2000a) Zhang et al. (1999) Zhang et al. (1997) Moon et al. (1998) Moon et al. (1996) Zhang et al. (1998) Ikeda et al. (1996)

Revised from Watanabe et al. (2000a)

total Cd absorbed, about 80–90% and 0.05% of the total ingested Cd is excreted in the faeces and urine, respectively. Most dietary Cd is bound to metallothionein and is absorbed intact into the circulation, and, in animals, kidney and liver Cd accounts for 50–70% of the total Cd with kidney having a higher Cd concentration than the liver. Other organs such as the pancreas, spleen, heart, brain, and testis, together with muscle and fat accumulate small amounts of Cd. In blood, Cd is associated with albumin-like protein, which is transported to the kidney where it is filtered through the glomerulus and reabsorbed by the proximal tubules (Friberg et al., 1985). In animals, the effects of prolonged exposure to abnormal levels of Cd include testicular necrosis, placenta destruction, abortion, teratogenic malformations, renal damage, osteomalacia, immunosuppression, pulmonary edema, and emphysema. In humans, severe Cd exposure can result in emphysema, bronchitis, ulceration of nasal mucosa, renal dysfunction, liver necrosis and anemia, hypertension, skeletal deformities, prostrate and lung cancer, and teratogenesis (Chowdhury and Chandra, 1987).

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Different countries have set guidelines on the maximum permissible levels for Cd in various meat products. In a survey of cattle, pigs, and sheep in New Zealand, the mean Cd concentration in kidney cortex, liver, and muscle was <0.4, 0.1, and 0.05 mg Cd kg−1 wet weight, respectively (Solly et al., 1981). These values are well below both the concentration measured in other countries and the maximum permissible concentration stipulated by many countries. However, Roberts et al. (1994) indicated that between 1988 and 1991, some 22–28% of sheep and 14–20% of cattle had kidney Cd contents greater than the permissible level of 1 mg Cd kg−1. In general, older animals had higher kidney Cd contents as these animals had longer exposure to Cd in their environment and, hence, greater opportunity to consume and retain Cd (Petterson et al., 1991). However, the rate of Cd increase decreases with age (Lee et al., 1996); for example, at 6 months of age, sheep retain about 0.1% of the Cd intake in the kidneys, with this value decreasing to 0.04% or less with increasing age (Loganathan et al., 2008). Feral deer and sheep isolated from human intervention also showed agerelated retention of Cd in kidney tissue as a result of exposure to naturally occurring Cd in the environment. Cadmium in cow and goat milk (0.002–0.006  µg  kg−1) (Rayment, 1988) and sheep and cattle muscle (0.01–0.03 mg kg−1 wet weight) were reported to be very low. Cadmium measured in carcass meat in New Zealand (0.02–0.03 mg kg−1 wet weight) (Roberts and Longhurst, 2002) and Australia (median, 0.01 mg kg−1 wet weight) (Langlands et al., 1988) were also significantly lower than the permissible level of 0.05  mg  kg−1 fresh tissue (kidneys, liver, muscle) in these countries (Loganathan et al., 2008). Because the accumulation of Cd in kidneys was high, whereas the amounts in milk and meat were low, Loganathan et al. (2008) suggested that the grazing animal can be viewed as a “filter” capable of ensuring that the amounts of Cd entering the food chain are low. They also conveyed that this was similar to the case, in which Cd-rich bran and roots were removed from processed grains and edible parts of plants, respectively, and then less Cd entered the food chain (Sauerbeck, 1992). The average annual consumption of red meat per New Zealander (70 kg adult) is estimated to be 80 kg. Based on a median Cd content of 0.01 mg kg−1 wet weight, the annual intake of Cd is expected to be 0.8 mg, which is equivalent to a daily intake of 2.19  µg Cd. This is considerably less than the maximum safe level of 70 µg Cd day−1 (Ledgard et al., 2010; Loganathan et al., 2008) from all sources in the diet. Although offal, such as liver and kidney, contains more Cd than the other meat, the small quantity

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of offal consumed by the New Zealand population is unlikely to contribute significantly to the total Cd intake. Also, as most meat-producing animals are slaughtered at a young age, this ensures that Cd concentrations of their edible offal products are within permissible limits (Loganathan et al., 2008). However, in sections of population and in pets where offal forms a larger portion of the diet, Cd intake from this source could be a major concern (Loganathan et al., 1999). The “population critical concentration” of Cd at which 10% of the human population would exhibit signs of renal impairment on the kidney is about 160 mg Cd kg−1 wet weight (Friberg, 1984). From the above calculations, it can be shown that it is unlikely that Cd input from meat exceeds the critical concentration in the life time of a New Zealander. However, Loganathan et al. (2008) proposed that computer-based models are required to identify farming systems that present a high risk of Cd concentrations in edible offal exceeding the permissible levels and livestock at risk of chronic toxicity. They recommended that a decision support model of this kind may be useful in developing management strategies capable of reducing Cd accumulation in animals. They stated that although preliminary empirical models have been developed for Cd accumulation in sheep grazing on New Zealand pastures (Loganathan et  al., 1999), further development of these models is required for their wider applicability.

5. RISK MANAGEMENT OF CADMIUM IN RICE ECOSYSTEMS Among heavy metals, Cd has been recognized as one of the most detrimental because Cd uptake is known to have caused itai-itai disease. The Codex Alimentarius Commission, set up by the UN Food and Agriculture Organization and the World Health Organization, has set standards 0.4 mg Cd kg−1 as the maximum permissible concentration of Cd in polished rice (Codex, 2006). Therefore, evaluation of the risk of Cd uptake is needed, as well as minimization of that risk by either decreasing soil Cd contamination or reducing Cd bioavailability to plants, which will improve food safety and safeguard human health. In this section, we describe appropriate methods to minimize Cd contamination in rice grains, including (1) decreasing Cd inputs to rice soils through the use of low Cd-­ containing P fertilizers, (2) water management to reduce bioavailability of soil Cd to rice plants, (3) replacement of contaminated soil with nonpolluted soil, (4) selection and breeding of low Cd-accumulating rice cultivar,

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(5) phytoremediation of the polluted soil by rice and other promising crops, and (6) chemical remediation of Cd-contaminated soil by soil washing with chemicals such as iron salts.

5.1. Decreasing Cd Inputs to Rice Soils The main sources of Cd inputs to rice soil include P fertilizers, biosolids, and mine tailings (Table 4.5). Cadmium input through P fertilizers can be reduced by either selective use of PRs with low Cd or treating the PRs to remove Cd. Superphosphate fertilizer manufacturers in many countries including New Zealand and Australia are introducing voluntary controls on the Cd content of P fertilizers. For example, the fertilizer industry in New Zealand achieved its objective of lowering the Cd content in P fertilizers from 340 mg Cd kg−1 P in the 1990s to 280 mg Cd kg−1 P by the year 2000 (Bolan et al., 2003a; Rys, 2011). The Cd content as determined by the PR source is the most difficult to control because supplies of PRs with low Cd contents are limited and sources with higher Cd contents continue to be used in many countries for practical reasons. A number of PRs (e.g. Jordan (El Hassa) PR and Morocco (Khouribga) PR) are low in Cd, and these can be used for the manufacture of superphosphates. Alternatively, since Cd has a low boiling point (BP = 767 °C), it can be removed by calcining the PRs. Phosphoric acid used in the food industry is manufactured mostly only after the removal of Cd through calcination of the PRs. Calcination of PRs may not be a likely option in the fertilizer industry because it is expensive and calcination decreases the reactivity of PRs making them unsuitable for direct application as a source of P (Ando, 1987). Chien et al. (2009) mentioned in a recent review that if a water-soluble P (WSP) fertilizer contains a high Cd content, granulation of WSP fertilizer with potassium chloride (KCl) may result in a higher Cd uptake by crops compared to the same, but bulk-blended PK fertilizer. They suggested that a possible explanation would be that in granulated PK fertilizers, KCl- and Cd-containing P fertilizers are in the same granule and thus are in close contact, thereby increasing the possibility of forming readily bioavailable CdCl20 and CdCl1+ complexes. They also added that it would be less likely that the complexes would form when KCl- and Cd-containing P granules are physically separated in bulk-blended PK fertilizers. The above hypothesis was tested and confirmed by Chien et al. (2003) in a preliminary greenhouse study using upland rice and soybean (Glycine max (L.) Merr.). In their study, all P and K sources produced by either granulation or bulk blending had the same granule size (1.68–3.36 mm diameter).

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1.8

28 Grain Straw

a

27

1.4 b

1.2

26 a

1.0

25 0.8 0.6

24 a

0.4 0.2

Grain yield (g pot–1)

Cd concentration (mg kg–1)

1.6

a

a

c

a b

0.0

GL (SSP + KCl)

BB (SSP + KCl)

23

22

BB (MCP + KCl)

Figure 4.2  Grain yield of upland rice and Cd concentrations in rice grain and straw. GL, granulated; SSP, single super phosphate; BB, bulk-blended; MCP, mono calcium phosphate (Chien et al., 2003). Means followed by the same letter within the treatments are not significantly different at p < 0.05.

The results showed that the agronomic effectiveness in increasing crop yield was the same with Cd-containing SSP and the reagent-grade monocalcium phosphate [(MCP) (0% Cd)], whether granulated or bulk blended with KCl. However, they noticed that concentrations of Cd in plant-tissue samples of all crops were much lower for MCP than for SSP. In all the planttissue samples, Cd concentrations obtained with granulated (SSP + KCl) fertilizers were higher than that with bulk-blended (SSP) + (KCl) fertilizers. Their results demonstrated that bulk blending of Cd-containing P fertilizers with KCl can reduce Cd uptake by crops compared to the same, but granulated, PK fertilizers (Fig. 4.2). Although PK sources, instead of NPK sources, were used in their study, they expected that inclusion of N will not affect the results, and, if proven true, the process of bulk blending, compared to granulation in decreasing Cd uptake, would also apply to NPK compound fertilizers.

5.2. Water Management to Reduce Cd Bioavailability Water management is a popular and cost-effective cultural practice for alleviating rice Cd contamination in Japan. Heavy-metal uptake by paddy rice, in particular, uptake of Cd, is considerably influenced by water management,

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hence, the redox state of the paddy soil.Table 4.14 shows the effect of water management on Cd content in rice grains. Cadmium absorption by rice was decreased drastically by continuous submergence of the paddy field after heading time. It is most likely that a considerable decrease in Cd absorption by rice under submerged conditions is due to a decrease in the Cd solubility because of the formation of carbonates (Khaokaew et  al., 2011) and/or CdS (de Livera et al., 2011; Iimura and Ito, 1978) as mentioned in Eqns (2) and (3) (Lindsay, 1979). CdCO3 (Octavite) + 2H+ = Cd2+ + CO2 (g) + H2O log K0 = 6.16 (2) CdS (Greenockite) = Cd2+ + S2− log K0 = −27.07



(3)

The former, Cd carbonate, is the primary form under alkali conditions (Khaokaew et al., 2011), while the latter, CdS, could be a dominant form under slightly acidic condition.The theoretical explanation based on physicochemical equation for the formation of CdS is summarized below. Flooding paddy field shifts Eh toward a reduced state (a sharp decrease in Eh), where the sulfate ion is reduced to the sulfide (Eqns (4) and (5)).

H2S (aq) + 4H2O = SO42− + 10H+ + 8e− (log K0 = −40.66)

(4)

Equation (4) can be rewritten by using Nernst’s equation with log K0;

Eh = 0.301–0.0739 pH + 0.00739 log (SO42−/H2S)

(5)

H2S is obtained from the dissociation of HS− and S2− with the following dissociation constants of K1 and K2. Table 4.14  Effect of Water Management on Cadmium Content in Rice Grains (Sakurai et al., 2005) Water Management after Heading Time Flooded Cd (mg kg−1)

Soil*

Soil A Soil B Soil C Soil D

Drained

Glasshouse Field test

Trace Trace 0.09 0.16

*Each soil is from different Cd-contaminated areas of Japan.

1.10 0.68 0.23 0.33

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K1 = (H+)(HS−)/(H2S) = 10–7.02

(6)



K2 = (H+)(S2−)/(HS−) = 10–12.9

(7)

Here, a total of the water-soluble sulfides is described: ΣH2S = (H2S) + (HS−) + (S2−)



(8)

Substituting the sulfide species in Eqn (8) using Eqns (6) and (7), Eqn (9) is obtained;

ΣH2S = (H2S) {1 + K1/(H+) + K1 K2/(H+)2}

(9)

At pH 7, Eqn (9) can be reduced to Eqn (10);

logΣH2S = 0.291 + log (H2S)

(10)

Finally, combining Eqns (4) and (10), Eh can be expressed by hydrogen sulfide and sulfate ion (Eqn (11)): Eh = −0.215 + 0.00739 log (SO42−/H2S)



(11)

or Cd extraction rate

(SO42–)/((SO42–) + (ΣH2S))

Assuming that dominant sulfur species in soil solution are H2S, HS−, S2− and SO42−, the change in the relationship between {(SO42−)/((SO42−)+(ΣH2S))} at pH 7 and Eh, is expressed by a solid line in Fig. 4.3 (Makino, 2002). The dashed line shows the relationship in a similar calculation at pH 6. On the other hand, the open circles show the measured extractable Cd with 1 M ammonium acetate from the soil in a water-submerged incubation test.The

Redox potential ( Eh )

Figure 4.3  Relationships between soil Eh and rate of Cd extraction or that of sulfate residue. The lines and open circles (pH 7, … pH 6, ○ measured value) are correspondent with the (SO42−)/{(SO42−) + (ΣH2S)} and Cd extraction rate of the vertical line in the figure. (Modified from Makino (2002) and Iimura and Ito (1978)).

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measured Cd data closely followed the calculated lines for the sulfide formation (Fig. 4.3), indicating that the Cd extraction rate is rapidly decreased with an increase in the ratio of ΣH2S to total S. Thus, the sulfide ion will precipitate with the Cd ion as CdS, whose solubility product is very small and hardly soluble in water (Eqn (3)). Flooding from tillering to head formation during rice growth would be the most effective period to decrease the Cd content in rice grains. It is highly recommended to keep flooding the paddy fields as late as possible toward harvest time. However, the later the flooding, the more difficult it is for machine operation of harvest, so that we have to find a balance between lowering bioavailable Cd and the difficulty in operating machinery. Arao et al. (2009) investigated the effects of water management in rice paddy on levels of Cd and As in Japanese rice grains. Their results indicated that flooding treatment after heading was more effective than flooding treatment before heading in reducing rice grain Cd without a concomitant increase in total As levels.

5.3. Low Cd-Accumulating Rice Cultivars 5.3.1. Genotypic Variation in Grain Cd Concentration in Rice Selection and breeding of low Cd-accumulating cultivars is the most ­cost-effective and environmentally friendly method for reducing the risk of contamination from Cd in food (Grant et al., 2008). Natural variations in the concentrations of Cd among cultivars have been well documented in staple crops including rice. Figure 4.4 shows the genotypic variation in grain Cd concentration in 35 rice cultivars grown in two types of Cd-­polluted paddy soils (Arao and Ae, 2003; Arao and Ishikawa, 2006). Cadmium concentrations in brown rice ranged from 0.13 to 4.31 mg kg−1 in Fluvisols and 0.79–7.65  mg  kg−1 in Andosols. The ranking of rice cultivars was maintained across the soils, suggesting that Cd concentration of rice grains could be controlled by genetic factors rather than environmental conditions. Generally, Cd concentrations are higher in Indica-type rice varieties than in Japonica ones. The representative Japanese Japonica cultivars, Nipponbare, Koshihikari, and Sasanishiki, were categorized as the rice group with low grain Cd concentrations. The lowest grain Cd concentrations were found in two varieties, LAC23 and HU-LO-TAO.These rice varieties can be used as good materials to develop varieties with lower Cd levels than those in the varieties currently under cultivation. Indica rice variety, IR-8, which led to the green revolution in the Asia in 1960s, was categorized as the highest group in grain Cd concentrations. Milyang 23 and Habataki were produced from the common ancestor, IR-8, and high Cd concentrations of these Indica

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varieties must be inherited from IR-8. These high Cd-­accumulating rice varieties can be used as the “cleaning plants” of Cd-contaminated paddyfield soil (see Section 5.5).Thus, large differences in Cd accumulation among rice cultivars enable us to develop phytotechnologies, such as breeding of low Cd-accumulating varieties and phytoextraction of Cd by using high Cd-accumulating varieties for reducing the Cd levels in rice grains. 5.3.2. Physiological and Genetic Mechanisms Understanding of the physiological and genetic aspects underlying Cd transport in rice is important to control Cd transfer into grains. The level of Cd in rice grains may be influenced by any of several physiological processes: 1) root Cd uptake, 2) sequestration of Cd into root vacuoles, 3) transfer from roots to shoots via the xylem, 4) transfer from xylem to phloem, and 5) phloem transport into grains. Uraguchi et  al. (2009) characterized the physiological properties involved in the differences in shoot and grain Cd accumulation between the low Cd-accumulating variety Sasanishiki and high Cd-accumulating variety Habataki.The activity of root Cd uptake was higher for Sasanishiki than for Habataki. However, Cd levels of xylem sap were well correlated with the shoot-Cd concentration in the two cultivars.

Figure 4.4  Genotypic variation in grain Cd concentration in rice. Thirty-five rice cultivars were cultivated in a container filled with two types of Cd-polluted soils under upland conditions (soil A, soil B). Koshikikari, Nipponbare, Sananishiki and Hu-Lo-Tao are Japonica and IR-8, Habataki and Milyang 23 are Indica subspecies. LAC23 is tropical Japonica (Javanica) variety (modified data from Arao and Ishikawa (2006)). A, Fluvisols (0.5 mg Cd kg−1 of dry soil); B, Andosols (5.1 mg Cd kg−1 of dry soil). For color version of this figure, the reader is referred to the online version of this book.

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Cd concentration in shoots (mg kg–1 dry wt.)

A positive and strong correlation between Cd concentrations in the xylem sap and subsequent shoot-Cd accumulation was also observed in a world rice core collection consisting of 69 accessions, which covers the genetic diversity of almost 32,000 accessions of cultivated rice (Fig. 4.5).These findings suggest that root-to-shoot Cd translocation via the xylem is the major and common physiological process determining shoot-Cd accumulation among rice cultivars. Rice-phloem sap can be collected using cut stylets of the brown plant hopper (Kawabe et al., 1980). Using this method, Cd levels in the phloem sap were measured and it was found that more than 90% of Cd present in grains is translocated via phloem (Tanaka et  al., 2007). Moreover, Cd concentration of the phloem sap of LAC23, the variety with lowest Cd accumulation in grains, was significantly lower than that of Koshihikari, the Japanese elite variety, despite similar levels of Cd in xylem sap in these cultivars (Kato et al., 2010). Thus, differences in grain Cd concentrations in rice cultivars may be in part explained by a different ability of phloem to transport Cd to grains. It is necessary to understand the genetic aspects of Cd accumulation in order to devise a breeding plan for reducing Cd levels in rice grains. Quantitative trait loci (QTL) analysis is a powerful tool for understanding the genetic control underlying agronomic and physiological traits in

60 50

40 30 20

Extra high Cd accumulating rice cultivars

10 0

0

200

400

600

800 –1

Cd concentration in xylem sap (µg L )

Figure 4.5  Relationship between Cd concentration in shoots and that of xylem sap in diverse rice germplasms (Uraguchi et al., 2009 with permission).

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rice (­Yamamoto et al., 2009). Using backcross inbred lines and advanced-­ backcross progenies derived from a cross between the low-Cd cultivar Sasanishiki and the high-Cd cultivar Habataki, Ishikawa et  al. (2010) reported a major-effect QTL (named as qGCd7) controlling Cd concentration in rice grains without affecting concentrations of essential trace metals (Cu, Fe, Mn, and Zn), and it is located on the short-arm chromosome 7. Moreover, this QTL had no significant effect on important rice agronomic traits, such as grain yield, grain weight, and days to heading. Using other mapping populations derived from crosses between low-Cd Japanese rice cultivars and high-Cd Indica ones, the QTL with a major effect related to Cd-translocating ability from roots to shoots at the seedling stage has also been detected on the short arm of chromosome 7 (Tezuka et  al., 2010; Ueno et  al., 2009). It also has been revealed that OsHMA3, a P1B-type ATPase, is a gene that controls root-to-shoot Cd translocation (Miyadate et al., 2011; Ueno et al., 2010). Functional analyses of the OsHMA3 gene in yeast showed that low-Cd cultivars contain a functional version of this gene, which is involved in Cd storage in root vacuoles. The high-Cd cultivars have lost this function; consequently, a much higher amount of Cd was loaded into the xylem. Overexpression of the functional OsHMA3 gene from the low-Cd cultivar drastically decreased Cd accumulation, not only in shoots but also grains in rice. Thus, this gene can be used to develop the phytotechnologies for controlling Cd accumulation in rice. 5.3.3. Breeding of Low Cd-Accumulating Cultivars Although Japanese rice cultivars are categorized as the low-Cd group and possess the functional OsHMA3 gene, some Cd-contaminated areas in Japan produce rice grains that exceed the maximum allowable limit of Cd. Thus, a breeding program has been initiated to produce rice varieties that have lower grain Cd concentration than the elite cultivars currently grown in Japan (Yamaguchi, 2006). LAC23, the tropical Japonica rice variety, was selected as a donor of the low-Cd trait because this variety has lower grain Cd concentration than other Japanese rice cultivars. LAC23 is not a practical variety in Japan because of late heading, long culms, long grains, and low yields. So crossing was undertaken with the Japanese rice cultivar Fukuhibiki, which has a good plant shape and offers stable high yields. This way, one could develop lines with low Cd concentrations but also with improved cultivation characteristics. By analysis across three to five self-fertilized generations (F3–F5), five promising lines were selected which, in comparison with Fukuhibiki and the elite cultivar Hitomebore, had 40–50% lower Cd

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Figure 4.6  Cadmium concentration in brown rice of newly developed lines. Five promising lines (named as Ukei1118-Ukei1122) were developed from a cross between LAC23 and Fukuhibiki at the National Agricultural Research Center for Tohoku region in Japan. Hitomebore is one of the popular cultivars in Japan.

concentrations in brown rice (Fig. 4.6), headed sooner than LAC23, and became comparatively shorter in plant height (Plate 4.2). These five lines were assigned the local numbers Ukei1118 through Ukei1122 based on the place where they were raised (National Agricultural Research Center for Tohoku Region, Daisen City, Akita Prefecture, Japan). For other trace metal concentrations such as Cu, Fe, Mn, and Zn, newly developed lines were nearly equal to those of Fukuhibiki and Hitomebore. Thus, it was possible to develop lines in which only the concentrations of Cd were reduced in the brown rice. However, further improvement should be done to incorporate high grain yield and good taste into promising lines. To develop practical low-Cd cultivars efficiently, attempts are being made to identify the QTL for the low-Cd trait controlled by LAC23 allele and to develop the DNA marker linked to the QTL for the screening process.

5.4. Soil Dressing Soil dressing is simple and is one of the most widely used techniques for heavily contaminated sites (Vangronsveld and Cunningham, 1998).This method has been adopted as a primary counter measure for Cd contamination in

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Plate 4.2  Plant shapes of the low-Cd line (Ukei1120) and parental cultivars (LAC23 and Fukuhibiki). Ukei1120 (one of the newly developed low-Cd lines) ripens sooner than LAC23 and has shorter culm length. For color version of this figure, the reader is referred to the online version of this book.

agricultural soils under the Agricultural Land Soil Pollution Prevention Law in Japan. Local managers, who are responsible for contamination prevention, prefer this technique over other countermeasures because of its low risk of failure, its predictable time frame, and because it leaves the site in a relatively pristine condition. There are several methods to amend the polluted soils by soil dressing (Yamada, 2007). 5.4.1. Simple Soil Dressing Unpolluted soils are placed on the top of the polluted soil (Fig. 4.7). Since the paddy fields amended by this method are raised by 20–30 cm, preparation of agricultural canals and agricultural roads and rezoning of paddy fields are needed with the application of this method. 5.4.2. Soil Removal Followed by New Soil Dressing Polluted surface soils are removed and discarded outside of paddy fields. Then, the infertile subsurface soils are covered with unpolluted soil. The

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depth of polluted soil removal is determined based on the degree of soil pollution and plant root elongation. 5.4.3. I n situ Placement of Polluted Soils First, the polluted surface soil is removed; then the subsoil is also temporarily removed to secure the place to bury the polluted surface soil (Fig. 4.8).

Figure 4.7  Simple soil dressing. For color version of this figure, the reader is referred to the online version of this book. (Modified from Yamada (2007)). Unpolluted soil Surface soil

Hardpan

(Unpolluted subsoil)

Subsurface soil

Subsoil

Removal of polluted Before works

Burying of polluted

Development of hardpan

surface soil

and placement of subsoil

surface soil and excavation of subsurface and subsoil

Figure 4.8  In situ placement of polluted soils (Yamada, 2007 with permission). For color version of this figure, the reader is referred to the online version of this book.

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The polluted surface soil is then buried into subsoil layer. After that, the part of removed subsoil is returned on top of the buried surface soil. This is followed by placement of new unpolluted surface soil on the top of the returned subsoil. However, it is difficult to apply this method to paddy fields where the subsoil layer soil is also polluted and/or the water table is high. This method is applied to paddy fields where the subsoil is not polluted. According to several follow-up surveys, soil dressing is an effective and reliable practice to decrease the Cd content in rice grains, when the newly added unpolluted soil layer is 20–30  cm thick. However, this practice is costly and becomes increasingly difficult for implementation because of the scarcity of suitable uncontaminated soils. Moreover, this method only postpones the problem to a later date and in authors opinion is not the preferred option.

5.5. Phytoremediation 5.5.1. Necessary Conditions for Phytoextraction Phytoextraction using hyperaccumulator plants has been proposed as a promising, environmentally friendly, low-cost technology for decreasing the heavy metal content of contaminated soils and has emerged as an alternative to the engineering-based methods (Ebbs et al., 1997; McGrath et al., 2002; Robinson et al., 2006). Hyperaccumulator plants can accumulate pollutants at high concentrations in their shoots and can grow in soils containing high concentrations of metals (Ebbs et al., 1997). Chaney et al. (2004) reported that some ecotypes of Thlaspi caerulescens in southern France showed high potential as a phytoextraction technology with low cost for soil Cd remediation. However, T. caerulescens may not be suitable for many large-scale phytoextraction projects because the plants are small and grow slowly, making them difficult to harvest mechanically (Ebbs et al., 1997). Cadmium-uptake efficiency of T. caerulescens in soils with relatively low levels of Cd pollution may be possible but it may not be effective in soils with more severe pollution (Brown et al., 1995). In addition, culturing these hyperaccumulator species could be hampered by their susceptibility to certain diseases. For example, McGrath et al. (2000) reported that several Thlaspi species are infected by diseases whose development was favored by prevailing humid and warm weather conditions. Because the typical weather conditions of the Asian Monsoon summer are humid and warm, it may be difficult to introduce these species into the Asian Monsoon’s paddy fields contaminated with low concentrations of Cd. To maximize the efficiency of phytoextraction, it is important to select a phytoextraction plant with high Cd-accumulating

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ability that is also compatible with mechanized cultivation techniques and local weather conditions. Such a plant may yield more immediate practical results than selection based solely on high tolerance to Cd. Several phytoextraction studies have tested nonhyperaccumulator highbiomass plants such as Indian mustard (Brassica juncea L.) (Ebbs et al., 1997; Nanda Kumar et al., 1995), tobacco (Nicotiana tabacum L.) (Mench et al., 1989), industrial hemp (Cannabis sativa L.) (Linger et al., 2002), flax (Linum usitatissimum L.) (Angelova et al., 2004), vetiver grass (Vetiveria zizanioides) (Chen et al., 2000), poplar (Populus spp.) (Laureysens et al., 2005), and willow (Salix spp.) (Hammer et  al., 2003). These plants can be cultivated in agricultural fields in Japan. However, rice is the biggest crop in Japan, and its cultivation system is well established and highly mechanized. The use of agricultural species adapted to growing conditions of paddy fields may, therefore, be a better alternative. 5.5.2. Plant Selection for Phytoextraction Rice, soybean, and maize (Zea mays L.) are the major summer crops grown in paddy fields and in upland fields (fields under aerobic soil conditions) that have been converted from paddies in Japan. However, the study of phytoextraction using rice and soybean has not yet been examined. Rice (cv. Nipponbare and Milyang 23), soybean (cv. Enrei and Suzuyutaka), and maize (cv. Gold Dent) were grown on one Andosol and two Fluvisols with a low concentration of Cd pollution ranging from 0.83 to 4.29 mg Cd kg−1, during 60 days in a greenhouse (Murakami et al., 2007). Shoot-Cd uptake was as follows: Gold Dent < Enrei and ­Nipponbare < Suzuyutaka and Milyang 23 (Fig. 4.9). Several soil-Cd ­fractions including exchangeable, inorganically bound, and organically bound decreased after harvesting of Milyang 23 (Fig. 4.10). Milyang 23 accumulated 10–15% of the total soil Cd in its shoot. These values were much higher than those reported for B. juncea (0.09%) and T. caerulescens (0.06%) grown on soil containing 40 mg kg−1 of total Cd for 6 weeks (Ebbs et al., 1997). The Milyang 23 rice is, thus, promising for phytoextraction of Cd from paddy soils with low pollution of Cd under aerobic soil conditions. 5.5.3. Phytoextraction by High Cd-accumulating Rice Soybean is the major summer crop grown in Japanese upland rice fields. The Codex Alimentarius Commission set maximum levels for Cd in wheat, potato (Solanum tuberosum L.), many vegetables (Codex, 2005), and polished rice (Codex, 2006). The commission discontinued work on developing

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a maximum level for Cd in soybeans, which it considered was not a major contributor to Cd intake (Codex, 2004). However, soybean, via tofu, natto, and soy sauce, is the main source of dietary intake of Cd in Japan (Arao et al., 2003). Thus, decreasing the Cd content of soybean seeds is extremely important. In an earlier study, Murakami et al. (2007) selected “Milyang 23” rice as a promising cultivar for phytoextraction of Cd in paddy soils with low-to-moderate pollution. However, its effect on the Cd content of subsequently grown soybean seeds has not been reported. Maize

Shoot-Cd uptake (µg pot-1)

350

300

b

250

Soybean (Suzuyutaka)

Rice (Milyang 23)

b b

b

200 150

100 50

a

0

b b

a

Andosol

a

Fluvisol 1

Fluvisol 2

Cd concentration in soil (mg kg–1)

Figure 4.9  Shoot-Cd uptake by maize, soybean, and rice. Error bars represent the standard error (n = 4). Means in the each soil followed by the same letter are not significantly different at p  <  0.05 based on Bonferroni’s multiple-comparison test. (Modified from Murakami et al. (2007)). 3.0

2.5 2.0 1.5

1.0 0.5 0.0

a a b b a No plant, fertilizer

a a

Res

a

Ox

a

Ino

b Maize

Org Ex

a a c

a a c

b

c

b

c

Soybean (Suzuyutaka)

Rice (Milyang 23)

Figure 4.10  Cadmium concentrations in five fractions of Fluvisols, which were remedied by three kinds of Cd accumulator plants. Ex, exchangeable fraction; Inorg, inorganically bound fraction; Org, organically bound fraction; Ox, oxide-occluded fraction; Res, residual fraction. Error bars represent the standard error (n = 4). Means in the each Cd fraction followed by the same letter are not significantly different at p < 0.05 based on Bonferroni’s multiple-comparison test. (Modified from Murakami et al. (2007)).

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To evaluate the effect of phytoextraction by rice on the seed Cd content of soybean grown subsequently, Murakami et  al. (2008) grew Milyang 23, a high Cd-accumulating rice cultivar, followed by soybeans in three paddy soils contaminated with moderate Cd concentrations (2.50– 4.27 mg Cd kg−1). The rice accumulated 7–14% of the total soil Cd in its shoots, and decreased several Cd fractions in soil by 8–42% and the total soil Cd by 10–19% (Fig. 4.11). Indian mustard (Brassica juncea (L.)) and the hyperaccumulator T. caerulescens accumulated 0.09 and 0.06% of the total soil Cd (40 mg kg−1) when grown for 6 weeks in pots (Ebbs et al., 1997). Nicotiana rustica L. accumulated 6% and N. tabacum L. accumulated 20% of the total soil Cd (5.44  mg  kg−1) when grown for 8  weeks in containers (Mench et al., 1989).Thus, Milyang 23 has the potential to phytoextract soil Cd with a similar efficiency as those Nicotiana species. The soybean seed Cd contents were 24–46% less than those grown on control soils (nonphytoextracted). Phytoextraction by Milyang 23 rice is, thus, a promising remediation method for reducing seed-Cd contents of soybeans grown on paddy soils under aerobic soil conditions. Previous research has shown that the Cd concentration in rice shoots grown under flooded (reducing) soil conditions may be low because Cd solubility under these conditions is lower than under oxidizing conditions

Figure 4.11  Cadmium concentrations in five fractions of control (nonphytoextracted) and phytoextracted soils (Andosol, Fluvisol 1 and Fluvisol 2) by Milyang 23. Ex, exchangeable fraction; Inorg, inorganically bound fraction; Org, organically bound fraction; Ox, oxide-occluded fraction; Res, residual fraction. Error bars represent the standard error (n = 3). **p < 0.01, *p < 0.05 (t-test). (Modified from Murakami et al. (2008)).

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(de Livera et al., 2011; Kabata-Pendias and Pendias, 2001), as mentioned in Section 5.2. However, the dry weight (DW) of rice shoots grown under flooded soil conditions is higher than that under oxidizing soil conditions during tillering (from transplanting to 30  days before panicle initiation) (Takahashi, 1974). Because shoot-Cd uptake by rice plants equals the product of DW and the Cd concentration of the rice shoots, maximizing shootCd uptake requires management practices that enhance both DW and Cd uptake by the rice shoot. Therefore, Murakami et al. (2009) undertook a field experiment in which the soils of all subplots were maintained under flooded conditions during tillering in order to maximize the DW of the rice shoots, and then drained them and kept them under oxidizing conditions until harvest to maximize Cd accumulation by the rice shoots (“without irrigation after drainage”) (Fig. 4.12). The total shoot-Cd uptake by the Indica Chokoukoku grown for 2  years (883  g  ha−1) was higher than that by the 3-year grown Indica Moretsu (869  g  ha−1), Indica–Japonica Milyang 23 (638  g  ha−1), and Indica IR8 (532  g  ha−1) (Fig. 4.13). This 2-year shoot-Cd uptake by Indica Chokoukoku from soil containing 1.63  mg  kg−1 of total Cd was higher than the uptake by the hyperaccumulator T. caerulescens (540  g  ha−1 after 3  years of cultivation in soil with a total Cd content of 2.8  mg  kg−1) (Hammer and Keller, 2003), by willow, Salix viminalis (170  g  ha−1 after 5  years of cultivation in soil with a total Cd content of 2.5  mg  kg−1) (Hammer et  al., 2003), and by poplar (Populus) clone Balsam Spire (57  g  ha−1 after 2  years of cultivation in soil with a total Cd content of 0.75  mg  kg−1) (Laureysens et  al., 2005). In contrast, Cd uptake by the residual roots of the Indica Chokoukoku was lower than that of the other Indica and Indica–Japonica rice cultivars

MD, Mid-season drainage;II, Intermittent irrigation; IF, Irrigation at flowering stage; r, Reduced; o, Oxidized

Figure 4.12  Water management during rice cultivation for normal (intermitted irrigation, upper) and for phytoextraction (without irrigation after drainage, lower). (Modified from Murakami et al. (2009)).

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Chokoukoku (I) Milyang 23 (IJ) Moretsu (I) IR8 (I) Akitakomachi (J)

b b b a

a

ab ab b

a c b c

b a b c –200

0

200

400

600

Shoot and root Cd uptakes by rice (g ha–1)

Figure 4.13  Shoot- and root-Cd uptakes by Indica-type rice cultivars capable of accumulating Cd at high levels and shoot-Cd uptake by Japonica food rice cultivar. Means in the same year (shoot or root) labeled with the same letter do not differ significantly (p  <  0.05, Tukey–Kramer’s HSD test). Error bars represent the standard error (n  =  2). J, Japonica; I, Indica; IJ, Indica–Japonica. Shoots were harvested in mid-October from 2004 to 2006. Residual roots were sampled in early May in 2007.  (Modified from Murakami et al. (2009)).

(Fig. 4.13). Cadmium in the residual roots may be released gradually into the soil as the roots are decomposed by soil organisms. Because phytoextraction involves harvesting of plant shoots that have taken up toxic elements from the soil and removing harvestable material from contaminated fields, plants such as the Indica Chokoukoku, with high shoot-Cd uptake and low root-Cd uptake, are ideal for phytoextraction. The shoot DWs of the four Indica and Indica–Japonica rice cultivars did not decrease, even after two or three continuous cultivations without irrigation after drainage, indicating that growth damage from continuous cultivation and the presence of toxic metals in the soil did not occur. This characteristic of rice is also useful for phytoextraction. The exchangeable, inorganically bound, organically bound, and total soil-Cd concentrations were lowest in the Indica Chokoukoku subplot, despite the fact that this cultivar was grown for only 2 years (Fig. 4.14).This suggests that this cultivar can take up Cd more efficiently than the other rice cultivars from the more resistant (inorganically and organically bound)

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Figure 4.14  Soil-Cd concentrations in the five fractions for each subplot, sampled before plowing in 2007. Ex, exchangeable fraction; Inorg, inorganically bound fraction; Org, organically bound fraction; Ox, oxide-occluded fraction; Res, residual fraction. Error bars represent the standard error (n = 2). Means in the same fraction that are followed by the same letter are not significantly different (p  <  0.05, Tukey–Kramer’s HSD test). No-p, no plant (control); IR, IR8; Mo, Moretsu; Mil, Milyang 23; Cho, Chokoukoku. I, Indica; IJ, Indica–Japonica. (Modified from Murakami et al. (2009)).

fractions, as well as from the more bioavailable (exchangeable) fraction.This uptake capability equaled that of the hyperaccumulator T. caerulescens when grown in pots (Hammer and Keller, 2002). The Cd uptake by the residual roots of the Indica Chokoukoku (29.5  g  ha−1) corresponded to only 0.02 mg kg−1 of soil Cd. Even allowing for the return of this root Cd to the soil by microbial decomposition, the total soil-Cd concentration in the Indica Chokoukoku subplot was 38% less than the mean value in the subplots with no plants (a reduction from 1.63–1.01 mg kg−1). This reduction in total soil-Cd concentration by the 2-year grown Indica Chokoukoku was higher than the reduction by 3-year grown hyperaccumulator T. caerulescens (by 15% of total soil Cd, assuming that this plant took up Cd from soil to a depth of 15 cm and with a bulk density of 0.85 Mg m−3) (Hammer and Keller, 2003). The Japonica food rice cultivar Yumesayaka grown after phytoextraction by the four Indica and Indica–Japonica rice cultivars and in the subplots without phytoextraction showed normal growth (Fig. 4.15). The average of the grain yields of Japonica Yumesayaka grown in the four subplots after phytoextraction and in the no plant subplot (5.1 Mg ha−1) was similar to that of

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Figure 4.15  Cadmium concentration in rice grain and grain yield of a Japonica food cultivar (Yumesayaka) grown after phytoremediation with Indica rice cultivars capable of accumulating Cd at high levels. Error bars represent the standard error (n = 2). Means in the each item that are followed by the same letter are not significantly different (p < 0.05, Tukey’s multiple-comparison test). No-p, no plant (control); IR, IR8; Mo, Moretsu; Mil, Milyang 23; Cho, Chokoukoku. (Modified from Murakami et al. (2009)).

Japonica food rice cultivars in Japan in 2007 (5.2 Mg ha−1) (MAFF, 2008). The grain-Cd concentrations of Japonica Yumesayaka grown after 2  years of phytoextraction with the Indica Chokoukoku were reduced by 47% (to 0.54 mg kg−1) compared to those of the same rice cultivar grown without phytoextraction (1.02 mg kg−1) (Fig. 4.15). Phytoextraction with the Indica rice Chokoukoku grown for 2  years without irrigation after drainage removed 883 g Cd ha−1, reduced the total soil-Cd content by 38%, and reduced the grain-Cd content in subsequently grown Japonica food rice by 47% without decreasing yield. The results suggest that phytoextraction with Indica Chokoukoku can remove Cd from paddy fields polluted with low-to-moderate levels of Cd and reduce the grain-Cd concentration of Japonica food rice cultivars to below the Codex standard within a reasonable time frame.This approach will help reduce the risk of Cd pollution of rice from paddy fields. Recently, phytoextraction has been criticized by several researchers because of the long period required for restoration, the difficulty of producing a high-biomass crop of the desired species, and the lack of knowledge of agronomic practices and management for phytoextraction (McGrath et al., 2006; Robinson et al., 2006).The research results of Murakami et al. (2009) should help to dispel these criticisms. The DW of, and Cd uptake by, the

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Indica rice Chokoukoku were higher than those in the hyperaccumulator T. caerulescens (Hammer and Keller, 2003). Paddy rice can be cultivated continuously (De Datta, 1981), and its cultivation system is well integrated and highly mechanized. The Indica rice Chokoukoku was managed by agricultural techniques familiar to farmers who grow Japonica food rice; it is, therefore, well suited to planting on a wide scale. The 2-year phytoextraction using Indica Chokoukoku without irrigation after drainage reduced the total soil-Cd concentration by 38%, and it reduced the Cd concentration in the grain of subsequently grown Japonica food rice by 47% without decreasing yield (Murakami et al., 2009). However, the grain-Cd concentration of the Japonica food rice was still above the Codex Alimentarius Commission’s international standard for the Cd content of rice grain (0.4 mg kg−1) (Codex, 2006). Although this study showed the shoot-Cd uptake by the Indica IR8 was lower than that by the Indica Chokoukoku, 3-year phytoextraction by Indica IR8 on a paddy field reduced the 0.1 mol L−1 ­HCl-extractable Cd concentration in soil from 0.48 to 0.33  mg  kg−1 and the Cd concentration in the grain of subsequently grown Japonica food rice to 0.11 mg kg−1 (Honma et al., 2009). Even if the rate of reduction of soil-Cd concentration by phytoextraction with Indica Chokoukoku were to become half of that in the first 2 years, an additional 2 years of phytoextraction by Chokoukoku would reduce the grain-Cd concentration of Japonica food rice Yumesayaka to below 0.4 mg kg−1. These results suggest that phytoextraction with the Indica rice cultivar Chokoukoku can remove Cd from paddy fields polluted with Cd at low-to-moderate levels and can reduce the grain-Cd concentration of the Japonica food rice cultivar Yumesayaka to below the Codex standard within a reasonable time frame. However, a potential hazard is in advertent use of the phytoextractor grain as a food for humans and domestic animals. For example, large numbers of people were poisoned in Iraq in the early 1970s when mercury-treated grain meant for seed was eaten in home-made bread; poisoning also occurred in the USA in 1969 when treated grain was fed to hogs, whose meat was subsequently eaten (Bakir et al., 1973; Curley et al., 1971). Phytoextraction by Indica rice cultivars capable of accumulating Cd at high levels will be applicable to the remediation of paddy fields in Monsoon Asia that have low-to-moderate levels of Cd pollution, provided that careful attention is paid to disposal of the high-Cd rice. Use of the phytoextraction techniques described here will help reduce the risk of Cd pollution of rice from paddy fields.

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5.6. Soil Washing Soil washing is conventionally performed ex situ using appropriate equipments, in which extracting reagents are used to remove hazardous metals from soil into aqueous solution (Elliott and Herzig, 1999). The in situ soil washing is usually called “soil flushing”; therefore, the word “soil washing” is consistently used in this section to avoid confusion. The soil-washing techniques offer a great advantage of high Cd-removal efficiency for contaminated soils. However, this technique has been considered to be difficult to apply directly to agricultural land because wastewater drained during the process of soil washing might pollute surrounding environments such as agricultural canals, neighboring agricultural fields, and groundwater. However, paddy fields possess an impervious hardpan just below the subsurface layer, which hinders vertical movement of water. The washed solution stays in the surface soil and does not penetrate into subsoil layers and groundwater. So, an in situ technology of soil washing should be utilized fully to take advantage of such unique characteristics of the paddy field. The in situ soil-washing method of paddy fields has to meet the following criteria (Makino et al., 2006, 2007): 1. U  se of washing chemicals with high efficiency of Cd removal but a minimal adverse impact on the paddy field and its surrounding environment. 2. C  ost-effective and environmentally sound in operation of the system. 3. S oil fertility of the paddy field and its crop growth are not greatly affected by the washing treatment or can be easily corrected by the application of agricultural materials. 4. T  he effect of washing can last for a reasonably long period. Makino and his team (Makino et al., 2006, 2007) have developed a new soil-washing practice combined with on-site wastewater treatment that completely satisfies the abovementioned four requirements, which is discussed in the following sections. 5.6.1. Selection of Washing Chemicals Chelating agents, neutral salts, and strong acids have been used for soil washing to solubilize metals (Davis, 2000). In particular, ethylenediamine tetraacetic acid (EDTA) has been commonly used due to its efficiency of Cd removal from contaminated soils (Abumaizar and Smith, 1999; Nakashima and Ono, 1979; Zeng et al., 2005). EDTA, however, is a persistent chemical and stays for a long time in the environment (Tandy et al., 2004). Some

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scientists, therefore, have used more biodegradable chelating agents instead of EDTA (Chang et al., 2005; Hong et al., 2002; Kantar and Honeyman, 2006; Mulligan et  al., 1999; Tandy et  al., 2004). In case biodegradable agents are used, however, the cost becomes relatively higher than the non-/ less-degradable counterparts. Makino et al. (2006) noticed that calcium chloride (CaCl2) is one of most appropriate soil-washing chemicals for Cd-contaminated paddy soils on the basis of Cd extraction efficiency, cost-effectiveness, and relatively low environmental impacts. The high efficiency of CaCl2 in the extraction of soil Cd was mainly attributable to the high selectivity of Ca for soil adsorption sites compared with monovalent cations, the concurrent lowering of the solution pH due to hydrolysis of exchangeable Al, and the formation of Cd–Cl complexes. Makino et al. (2006) also mentioned that hydrochloric acid (HCl), nitric acid (HNO3), and EDTA-2Na extracted more soil Cd than neutral salts (Figs 4.16 and 4.17). However, EDTA-2Na is difficult to use for practical purposes, because of its persistent nature in the environment and relatively high cost. Both strong acids cause serious soil acidification, which is a problem if the soils possess a low pH-buffering capacity. Iron (III) chloride (FeCl3) extracted nearly as much Cd as HCl, HNO3, and EDTA-2Na from the Nagano (Fluvaquents) and Toyama (Epiaquepts) soils and more Cd from the Hyogo soil (Fluvaquents) (Fig. 4.16) (Makino et  al., 2006). Iron is a major soil constituent and is less environmentally harmful than either of the other chemicals. In addition, FeCl3 is less expensive and easier to handle than both HCl and EDTA-2Na, so that FeCl3 was selected as a promising washing chemical. The Cd extraction capacity was compared with other metal salts to elucidate the mechanism of Cd extraction by FeCl3. The proportion of total soil Cd extracted by the washing chemicals (i.e. the Cd extraction efficiency) increased in the following order: Mn salts ≤ Zn salts << ferric Fe salts in all three soils, with efficiencies ranging from 4 to 41%, 8 to 44%, and 24 to 66%, respectively. The amount of Cd extracted was negatively correlated with the extraction pH, suggesting that extraction pH plays an important role in determining the Cd-extraction efficiency. When metal salts are added to soils, the dissociated metal cations may form hydroxide precipitates with the release of protons (H+) according to the following equations (hydrolysis):

MmAn = mMn+ + nAm−

(12)

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Nagano soil

Hydrochloric acid Nitric acid Citric acid Acetic acid EDTA-2Na Iron( ) chloride -0.13

1E-15

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Cd extracted (mg kg–1 oven-dry basis) Hyogo soil Hydrochloric acid Nitric acid Citric acid Acetic acid EDTA-2Na Iron( ) chloride 0

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3

Cd extracted (mg kg–1 oven-dry basis) Toyama soil Hydrochloric acid Nitric acid Citric acid Acetic acid EDTA-2Na Iron( ) chloride 0

0.1

0.2

0.3

0.4

0.5

Cd extracted (mg kg–1 oven-dry basis)

Figure 4.16  Effect of various chemicals (other than neutral salts) on the efficiency of soil Cd extraction; concentration of chemicals used for the Cd extraction (0.02 molc L−1 (▨) and 0.1 molc L−1 (□)). (Makino et al. (2008) with permission).

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Potassium acetate Potassium chloride

Nagano soil

Sodium acetate Sodium chloride Magnesium acetate Magnesium chloride Calcium acetate Calcium chloride

0

0.04

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Cd extracted (mg kg–1 oven-dry basis) Potassium acetate Potassium chloride

Hyogo soil

Sodium acetate Sodium chloride Magnesium acetate Magnesium chloride Calcium acetate Calcium chloride 0

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(Cd extracted (mg

0.09

0.12

kg–1 oven-dry

0.15

basis)

Toyama soil

Potassium acetate Potassium chloride Sodium acetate Sodium chloride Magnesium acetate Magnesium chloride Calcium acetate Calcium chloride 0

0.05

0.1

Cd extracted (mg

0.15

0.2

kg–1 oven-dry

0.25

basis)

Figure 4.17  Effect of various neutral salts on the efficiency of Cd extraction from three soils; concentration of the neutral salts used for the Cd extraction (0.02 molc L−1 (▨) and 0.1 molc L−1 (□)). (Makino et al. (2008) with permission).

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Mn+ + nH2O = M(OH)n+ + nH+−

(13)



Kom = [M(OH)n][H+]n/[Mn+][H2O]n

(14)

where MmAn denotes the metal salt, M, the metal cation (Fe, Zn, or Mn), and A, the anion (Cl−, NO3−, or SO42–); m and n represent the charge numbers of the anion and cation, respectively; Kom denotes the equilibrium constants (expressed in terms of activities) for metal Mn+ in Eqn (13), which correspond to 2.88 × 10−4, 3.31 × 10−13, and 6.46 × 10−16 for Fe3+, Zn2+, and Mn2+, respectively (Lindsay, 1979). The precipitation of the metal hydroxide (hydrolysis of the metal ion) generates H+ at a rate that depends on Kom, and these protons may decrease the extraction pH (Eqns (12)–(14)). Figure 4.18 illustrates the theoretical relationships between pH and activity of metal ions in the metal hydrolysis reactions at the equilibrium with soil iron (calculated using Eqn (14) and the Kom values). The pH of ferric hydroxide is around 2 (Fig. 4.18), which is much lower than the original pH of the three soils. Thus, the Fe hydrolysis is associated with a greater decrease in soil pH compared to the other two metals. This indicates that a driving force of the Cd extraction by FeCl3 is proton release, which results in a sharp decrease in soil pH. In another study, Cd was highly mobile under the oxidizing and acidic conditions seen in these soils (Kabata-Pendias, 2000). Heavymetal solubilization was greatly enhanced by acidification and, at pH 1.3, reached more than 80% of the total Cd content of the soil (Dube and ­Galvez-Cloutier, 2005). Results obtained in the study endorse the effectiveness of iron salts as the washing chemical to remove soil Cd. 10 Mn (OH)3

8

pH

Zn (OH)2

6

4 Fe (OH)3

2

0 0.000

0.002

0.004

0.006

0.008

0.010

Activitity of free metal ions

Figure 4.18  Diagram of pH and metal activity to precipitate metal hydroxides. (Makino et al. (2008) with permission).

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The Cd-extraction efficiency of metal chlorides was greater than that of the corresponding metal sulfates and nitrates in all soils. Extraction efficiency decreased in the following order: chlorides > nitrates ≈ sulfates, with values ranging from 41 to 75%, 14 to 63%, and 26 to 62%, respectively, in the Nagano soil. The results were similar for the other two soils. To examine the factors that resulted in the difference in the extraction efficiency between the metal salts, Makino et al. (2008) estimated the relative abundance of dissolved Cd species in 100 mmol L−1 iron salt solution by the Visual MINTEQ software (Gustafsson, 2004). Figure 4.19 indicates that Cd–Cl complexes such as CdCl+ and CdCl2 (aq) accounted for 80% of the total dissolved Cd in the Nagano soil at 100 mMc FeCl3, versus values of 33% for Fe2(SO4)3 and 9% for Fe(NO3)3. Similar trends were observed for the other metal salts and soils. Cadmium has a high capacity to form complexes with anions such as Cl−, SO42–, CO32–, PO43–, organic acids, and ­fulvic acid (Kunhikrishnan et  al., 2012; Traina, 1999). Doner (1978) reported that Cd was leached more rapidly in the presence of Cl− than in the presence of ClO4−. Sakurai and Huang (1996) showed that the rate of desorption of Cd from a montmorillonite was greater with KCl than with KNO3. Smolders and McLaughlin (1996) suggested that high concentrations of Cl− in saline soils might increase plant uptake of Cd either by enhancing mass transport of Cd or by enhancing uptake of the CdCl+ complex by plant roots. Accordingly, the formation of stable Cd–Cl complexes

Figure 4.19  Relative abundance of various Cd species in the extracts of the Nagano soil in the presence of the three iron compounds. The Cd species were calculated using the Visual MINTEQ software (Gustafsson, 2004) based on the dataset of cations, anions, pH and dissolved organic carbon values obtained for the extracts. (Makino et al. (2008) with permission).

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could inhibit resorption of the extracted Cd onto adsorption sites on the surface of the soil particles. This inhibition mechanism will improve the efficacy of extraction with FeCl3 compared to that with Fe2(SO4)3 and Fe(NO3)3, because the proportion of Cd complexes to the total dissolved Cd concentration is high in the extracts with chloride salts. 5.6.2. On-site Soil Washing (Soil Flushing) in Paddy Fields The soil-washing procedure developed by Makino et al. (2007) consisted of three steps: (1) chemical washing with appropriate chemical solutions, such as CaCl2 and FeCl3 to extract Cd from soils, (2) followed by water washing to eliminate the remaining chemicals, and (3) on-site treatment of the wastewater by a portable purification apparatus with a chelating material (Fig. 4.20). In a field study, a part of the paddy field was bound with plastic boards, which were partially buried on the edges of the paddy field so that the upper two-thirds of each board remained above the ground surface.This boundary provided containment for additional water and chemicals in the paddy field. The flushing chemical was applied to the bounded experimental field, followed by addition of agricultural water, creating a soil-solution ratio of 1:1.5–1:2.The soil suspension was mixed by a tilling machine until it turned into slurry. It is important to mix the soil suspension thoroughly enough because the structure of soil clods was maintained and the diffusion of washing chemical into the clods was likely a rate-controlling factor for the extraction

Figure 4.20  Conceptual diagram of on-site soil washing. For color version of this figure, the reader is referred to the online version of this book.

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of soil Cd. The soil suspension was allowed to rest after the mixing, and then the Cd-containing supernatant of the slurry was drained off as wastewater and sent to the wastewater treatment system (Makino et al., 2007). The Cd concentrations in the treated wastewater were below the ­Japanese environmental quality standard (0.01 mg Cd L−1), demonstrating that the in situ treatment system could treat the wastewater as expected.The Cl concentration was <500 mg L−1 after water washes; this concentration is the threshold value for healthy rice crops. The concentration of exchangeable Cd was significantly decreased and little changed after CaCl2 and FeCl3 washing, respectively, but the weakly acid-soluble Cd form decreased substantially in the FeCl3 washing. Although the exchangeable Cd increased with decreasing soil pH caused by the washing treatment, adjusting the pH to the initial pH by the addition of lime could decrease the exchangeable Cd concentration and maintain it at this level after the washing. The soil pH values were significantly decreased after the washing treatment. Although EC increased, it did not reach a level that would affect growth of rice plants. Exchangeable Mg and K decreased owing to the soil washing. The Mg and K deficiencies were corrected by the application of fertilizers to the washed soil, restoring the Mg and K concentration in soil during the growth period.Total carbon and total N concentrations changed little. Although the extraction pH became very acidic with an application of FeCl3, the amount of soil Al released was <1% of the total soil Al, indicating that the in situ soil treatment is unlikely to cause serious soil damage such as clay mineral destruction. Soil washing considerably decreased the Cd concentrations in the rice grain. The reduction rates of unpolished rice after CaCl2 and FeCl3 washings were around 40 and 80%, respectively. These results proved the ­efficiency and effectiveness of the soil-washing method for remediation of Cd-contaminated paddy fields.

5.7. Integrated Risk Management A number of challenging issues need to be taken into consideration when devising strategies to manage Cd contamination in rice ecosystem. These include the following: 1. M  ultisource of Cd contamination: Cd reaches rice ecosystems through various sources that include Cd-containing P fertilizers and organic amendments, drainage from mine tailings, and domestic and industrial wastewater.

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2. C  omplexity of Cd contamination: the severity and long-term persistence of Cd contamination in rice ecosystems are influenced by factors such as site hydrogeology, redox conditions, the flooding period, water use, and chemical form and speciation of Cd. 3. C  hanges in chemical speciation: Cd undergoes several biogeochemical transformation processes, especially under alternate wetting and drying conditions prevalent under rice cropping, thereby resulting in the release of Cd species which differ in their biogeochemical reactions, bioavailability, and biotoxicity. 4. M  agnitude of Cd contamination resulting from excessive use of irrigation water: Lowland rice consumes large quantities of water, resulting in a high uptake of Cd in the presence of Cd-rich wastewater irrigation. ­Similarly, high uptake of As has been found to occur with the use of As-rich irrigation water in rice cropping in Bangladesh and India (Mahimairaja et al., 2005). It is, therefore, important to formulate integrated risk-management strategies involving source avoidance, source reduction, and remediation of soil and water (Fig. 4.21). Source avoidance refers to avoiding the Cd-­ contaminated sources such as fertilizer, organic manure, irrigation water, mine tailings, and Cd-contaminated sites for rice cultivation. For example, Cd-free P fertilizers such as DAP can be used to supply both N and P for rice crops. Another strategy is source reduction, which refers to removing or stopping the source of Cd contamination and subsequent uptake by rice plants. Source reduction can easily be achieved when the contamination source is of anthropogenic origin such as mine sites or similar point sources. Unlike in the case of As contamination, in most regions, Cd contamination of rice ecosystems is largely of anthropogenic origin, and source reduction may be a feasible option to manage Cd contamination. Remediation of contaminated soil and water resources requires both short-term and long-term solutions for the Cd problem. Depending on the efficiency and cost-effectiveness of the system, a combination of technologies may be required. The potential technologies for remediation of Cd-contaminated soil and water resources in relation to rice cultivation are presented in Fig. 4.21. For example, in the case of soil, a number of technologies including soil dressing followed by soil washing, immobilization of Cd using liming materials, and phytoremediation are available to manage Cd contamination. In the case of irrigation water, immobilization of Cd following by filtration can be used to remove Cd (i.e. Cd stripping). More sophisticated stripping methods, which may require a series

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Cd in rice ecosystem

Source reduction Site

Input Low Cd fertilizers and manures

Avoiding mine influenced sites

Avoiding mine tailing drainage

Avoiding acidic Cd soils

Limiting industrial effluents

Mitigating tailings dispersal

Rice field

Soil Mobilization

Water

Rice plant

Immobilization

Soil washing

Cd immobilization

Immobilization

Low Cdaccumulated cultivar

Electro osmotic flow

Phytoremediation

Filtration

Flooding treatment post heading

Optimise N fertilizer form & rate

Soil dressing

Foliar silica sols

Molecular bondingTM stabilizer

Figure 4.21  An integrated approach to manage Cd in rice ecosystem.

of filtering–sorptive (precipitation) setups, are necessary in order to cope with the enormous volume of irrigation water required for rice cultivation. Cultural practices to manage Cd accumulation in rice grains include cultivation of low Cd-accumulating rice varieties and flood management after heading. Hence, a successful remediation scheme for Cd-contaminated rice ecosystems should aim for an integrated approach involving the possible combination of physical, chemical, and/or biological mechanisms. It is essential that the integration of remediation technologies should enhance efficiency, both technologically and economically, resulting in a reduction in the time required for achieving targeted levels of Cd in the

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rice grain. For example, phytoremediation is a promising new technology, which is relatively inexpensive, and proven effective in large-scale remediation of both soil and water resources. Further, it would also add “green” value (esthetic) to the environment. Integrating physical, chemical, and/or bioremedial measures with phytoremediation could enhance higher uptake of Cd by sacrificial plants, thereby achieving the removal of Cd from contaminated rice fields.

6. SUMMARY AND FUTURE RESEARCH NEEDS Cadmium (Cd) is one of the major toxic heavy metals reaching the food chain. In rice soils, Cd is derived mainly from the application of Cdcontaining P fertilizers and biosolids. Contamination also results from mine tailings and acid-mine drainage. Cadmium is accumulated in plants more readily than most other metals and can be translocated into edible parts before any signs of phytotoxicity. Rice is one of the most important crops grown for human consumption, and Cd accumulation in rice grains poses a potential health risk. The enrichment of Cd in paddy rice grain tends to occur during soil oxidation, which accompanies preharvest drainage of the flooded paddy. Several studies about Cd-contaminated soils are available but the guidelines established by individual countries worldwide to control the pollution of agricultural soils are not consistent and standardized. Scientific evidence from site-specific research, particularly long-term field trials involving all types of key conditions and factors, are necessary to understand the bioavailability of Cd in various soil types and to provide reliable parameters for health-based risk assessments. In particular, the collection of reliable databases on Cd concentrations in different conditions of soils, climates, rice cultivars, etc. must be given utmost importance. From these databases, a variety of useful information and methodologies can be developed toward achieving the ultimate goal of providing safe and high-quality rice, especially in the case of lowland soils. Unlike upland soils, paddy soils are flattened evenly for water control under flooding and huge amounts of water are irrigated during rice cultivation. The irrigation of contaminated water and the utilization of contaminated soils are considered to be the most important routes of Cd accumulation in paddy soil and rice plants. The best solution for mitigating Cd contamination in paddy soils and the rice plant is to remove the sources of Cd in the environment and to prevent Cd flow into paddy soils. Hence,

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further research needs to be done to determine the effectiveness in suppressing Cd availability when released from the sources, which include mining and industrial wastes and others.  A number of soil remediation techniques for Cd-contaminated soils have been developed but some of these techniques are not efficient in terms of time, cost, and environmental compatibility. In order to select the best and most practical technique for remediation of Cdcontaminated rice paddy soils, more investigations are needed. The biogeochemistry of Cd in rice ecosystems is complex and mostly determined by its chemical speciation resulting from chemical and biological transformations.The chemistry of soil and water (i.e. pH and Eh) plays a major role in Cd dynamics in paddy soils, which undergo changes in redox reactions during a growing season. Risk management of Cd-contaminated paddy soils and the associated flooded water is an important issue and a great challenge. Its success is necessary to minimize Cd accumulation in paddy rice grains that reach the food chain. A number of physical, chemical, and biological technologies involving immobilization, filtration, and phytoremediation have been developed to remediate Cd-contaminated paddy soils and the associated water. Conventional physical and chemical remedial measures usually are expensive but may prove highly effective. Field testing of some of these technologies has shown them to be successful in reducing Cd accumulation in rice grains. Low Cd-accumulating rice varieties can be used to minimize Cd reaching the food chain. Phytoremediation, which is relatively inexpensive, has been proven effective in the remediation of metal(loid)s-contaminated sites, including those with Cd. Nonedible, Cd-hyperaccumulating crops, like ornamental and fuel crops, may be suitable for phytoremediation through which the entry of Cd into the food chain could largely be avoided. Remediation of Cd-contaminated rice soils and Cd stripping from irrigation waters require an integrated approach involving a combination of physical, chemical, and biological technologies for successful and effective management of Cd-contaminated rice ecosystems. Future research is, therefore, needed along the following lines: • Elucidation of soil and water environmental factors (e.g. pH and Eh) that govern transformations of Cd both in upland and lowland rice ecosystems. • Examination of solid-phase and solution-phase speciation of Cd in soil and water using advanced spectroscopic-based techniques. • Identification of biochemical mechanisms involved in the accumulation of Cd in paddy rice grains.

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• Rhizosphere processes underpinning effective phytoremediation technologies for Cd removal from paddy soils. • In situ immobilization techniques in paddy soils using inexpensive contaminant-free industrial byproducts high in FeCl3. • Highly effective and expensive stripping methods for the removal of Cd in water supplies destined for irrigation.

ACKNOWLEDGMENTS The senior author thanks CRC CARE for providing funding (No 2-3-09-07/08) to undertake research on landfill site remediation. Drs Makino, Ishikawa and Murakami are grateful for the grant from the Ministry of Agriculture, Forestry and Fisheries of Japan (Research project for ensuring food safety from farm to table AC-1310, -1320); part of the review was derived from the abovementioned projects. The Postdoctoral fellowship program (PJ008650042012) with Dr Won-Il Kim at National Academy of Agricultural Science, Rural Development Administration, Republic of Korea, supported Dr Kunhikrishnan’s contribution.

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